Bioaccumulation Factors for Radionuclides in Freshwater Biota

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Contract No. W-7405-eng-26

Environmental Sciences Division

ULATIBN FACTORS FOR IN FRESHWATER BI

Henry A. Vanderpl oeg Dennis C. Parzyck ... William H. Milcox James R.. Kercher Stephen ‘1. Kaye

iii

...

This report analyzes over 200 carefully selected papers to provide concise data sets and methodology for estimation of bioaccumulation factors for tritium and isotopes of strontium, cesium, iodine, manganese, and cobal t in major biotic components of freshwater environments Bioaccumulation factors of different tissues are distinguished where significant differences occur. Since conditions in the laboratory are of ten unnatural in terms of chemi cal and ecological re1 ati onships

this review was restricted as far as possible to bioacciiniulation factors determined for natura? systems. Because bioaccumu ation factors were not available for some shorter-lived radionuc ides, a

methodology for converting bioaccurnul ation factors of stab e isotopes to those of shorter-lived radionuclides was derived and utilized. The bioaccumulation factor for a radionuclide in a given organism or tissue may exhibit wa’de variations among bodies of water that are related to differences in ambient concentrations of stable-element and

carrier-element analogues. To account for these variations simple models are presented that relate ~ioac~u~~~a~~o~factors to stable-

element and carrier-element concentrations in water. The effects of

physicochemical form and other factors in causing deviatfcmns from

these models are discussed. Bfoaccumulation factor data are examfned in the context of these models, and bioaccurnulation factor relations

for the selected radionuclides are presented.

S UMMA R Y

BIOA~C~~U~ATI~~FACTOR CONCEPTS The biOWXUMulatiOn factor for an organism or tissue i is’the -state ratio of radionuclide concentration in the organism or tissue to that in water:

where

BF(RIi = bioaccumulation factor for radionuclide R in organism or tissue i,

[RIi = radionuclide concentration (pCi/g fresh wejght) in

organism or tissue i and

[RIW = radionuclide concentration in water ($i/g)3 a constant. Bioaccumulation factors are used to p-edict radionuclide c~~c~~~r~t~~~~ in whole organisms or their tissues from knowledge of the ~ad~~nL~~li~~ concentration in water for chronic releases of radionuclides. Bjo- accumulation factors for radionuclid2s were related to ambient concentrations of their stable or novisotopic carrier element analogues according to three idealized patterns. The first pattern is that the bioaccumulation factor for radionuclide R in organism or tissue i, BF(RIi, is constant regfirdless of stable-element or carrier-element concentration :

The second pattern applies to an element that is homeostatically Vi maintained at a constant concentration in organism or tissue i:

Where

Xi = concentration of stable element in organism or tissue

i , a constant (ug/g fresh weight) , and

[CIkJ '= concentration of corresponding stable element in water

(vg/g 1 *

The third pattern applies to a radioisotope and its non-isotopic carrier element which is horneostatical ly maintained at a constant Concentration in i:

qi = discrimination coefficient,

Zi* = concentration of non-isotopic carrier element in organ-

ism or tissue i, a constant (Iig/g fresh weight), and

[C*3, = concentration of nsn-isotopic carrier element in

water (ug/y) "

The discrimination coefficient, qi , is the ratio ([R]i/lC*]i)/( [RIP// [C*Iw), where [C*], is the carrier element concentration in i. Equation (3) is often rewritten in the form: vii

Radionuclides exist in a wide variety of physicochemical forms in natural waters, and their different forms have different availabilities to the food chain, Sediments, too, may be source of radionuclides to biota, and sediment type can influence the availability. For those elements that are hsmeostatically maintained at constant concentrations in a giver organism, the concentration of stable element in the organism is independent of concentration of stable element in water or its availability in prey, sediment, and different physicochemical forms in the water. In contrast, differences in availability of radionucl ides in different sediment types and different physicochemical forms in water can 'lead to niarked deviations from the idealized patterns of Equats'ons (1) and (3)- lrlhen bioaccumulation factors were not available for radiunuclides they were estimated froni bioaccumwl ation factors of the corresponding stable elements. Owing to physical decay, the bioaccumulation factor of a short-lived radionuclide is much less than that of the stable element or longer-lived nuclides. 5iaaccumulation factors of shorter-

1 ived radionuclides were estimated from bioaccumulation factors sf the stable element according to the relation:

2 BF(R)i = r<.q 5FWi 7 where

k = elimination coefficient of C in i (day-'), A = radioactive decay constant (day-')3 and viii

BF(C) = bioaccumulation factor of corresponding stable elemerat in i.

B IOBCCUMULA'T ION FACTORS

Potassium is a non-isotopic carrier elerrretit For cesiiim because of the?'t- chemi sa7 simi 1 ari ties and the greater abundance of potassi iarn in water. Further, since potassium concentration is horncostatically maintained 3t constant concentrations in animals, the bioaceumulation factor of cesiuin -is given by Equation (3) or (4). Unlike this pattern for animals, potassium concentration of water has only a smal 1 effect on the cesi urn bi oacciiinul ation factor in a1 gae. The primary mode of accimulation of cesium and pstai;siurn in aquatic animals is Prom the food chajn. Absorptl'on efficiency of potassium and cesium from food is high. In animals the excretion coefficient of potassium is about. 3 times larger than the excretion coefficient of cesium. As a result, qi increases by a factor of about 3 with each trophic level. If potassiiern concentrations in the predators and prey are about equal, the cesium bioaccumulation factor increases by a factor of 3 with each trophic level.

Because cesi urn a's strmgly adsorbed by suspended particulate materials, especially clays, the proportion of cesium in the soluble phase decreases w'ith increasing suspended solids concentrations.

Potassium, too, is sorbed but to a much lesser degree. Since algae accumulate cesium, potassim, and other elements from the soluble phase, the availabilities to the food chain of cesium and of cesium ix

relative to potassium decrease with increasing suspended solids concentrations. Thus, discrimination coefficients and bioaccumulation factors decrease with increasing suspended solids concentrations. On the basis of data available, we recommend the bioaccumulation factors given in Table 1.

Stronti um

Cal ci um is a non- i so topi c carri er el emen t for stronti um because of their chemical similarities and the greater abundance of calcium in water. Further, since calcium concentration is homeostatically mas’ntained at constant concentrations in animals, the bioaccumulation

factor of strontium is given by EqLation (3) or (4). Unlike this pattern for animals, calcium concerttration of water has only a small effect on the strontium bioaccumulation factor in algae, a The primary mode of uptake of strontium and calcium in animals (as

well as plants) is from water. As a result, trophic level has little effect on the discrimination coefficient and the bioaccumulation factor of strontium. Further, the discrimination coefficient is

relatively independent of calcium concentration in water. Calcium and strontium concentrations in bones and shells are higher than in other tissues of animals. Strontium has physicochemical properties similar to calcium and,

like calcium, appears mainly as free ions in water. As a result of this and the fact that the discrimination coefficient is independent of calcium concentration in water, the discrimination coefficient varies little among sites. This implies that the product qici * X

Table 1. Recommended bioaccumulation factor relations for 1 ong-1 ived isotopes of cesi urna

...... __I.. .. .- Yaxsn/ Recommended Functional Tissue Envi ronrnent Bioaccumul ati on Group Factor Re1 ation , ......

Piscivorous b fishes A17 tissues C1 eay waters

Piscivorous d fishes A11 tissues Turbid waters

on -p i s c i vor ou s M fishes All tissues Clear waters Non-piscivorous fishes All tissues Turbid waters A1 gae Whol e A1 1 waters Aquatic vascul ar plants Whol e A1 1 waters 1 o3 Emergent vascular plants Whol e A1 1 waters Mol 1uscs She1 1 A1 1 waters Soft tissues All waters Invertebrates other than mol 1 uscs WhQl e All waters I o3

Amphibians All tissues A? 1 waters 1 o4 Waterfowl A?? tissues All waters 3x1 o3

I- aTo convert to bioaccumulation factors of 13%s use conversion factors in Table 3.1.1. bSuspended sol ids concentration less than 50 ppn.

C [Klw = stable potassium concentration of water in ppm. dSuspended sol ids concentration greater than 50 ppm. xi appearing in the numerator of Equation (3) may be treated as a constant and that a regression of In BF(Sr)i versus 1n[Calw, where [Ca], is calcium concentration in water, will have a slope near -1 and a high correlation coefficient.

On the basis of the data available, we recommend the strontium bioaccumulation factors given in Table 2 and shown in Figure 1.

Tritium

Tritium was included in this report because of a concern that the relative kinetics of tritium and protium resulting from tritium's heavier mass ("isotope effect") might lead to a preferential accumulation of tritium over protium. Experiments in aquatic systems indicate that this does not occur and that the bioaccumulation factor for tritium is less than or about equal to the bioaccumulation factor for protium, which has a bioaccumulation factor approximately equal lo 1. We recorn- mend that a bioaccumulation factor of 1 be used for all tissues of all aquatic biota (Table 3).

Iodine

Iodine is highly concentrated by the thyroid tissue of vertebrates. As a result, the bioaccumulation factor for iodine in the thyroid tissue of fishes is very high. Recomended bioaccumulation factors are given in Table 4.

Man gane s e

Manganese is homeostatically riaintained at constant concentrations in vertebrates and some invertebrates. Thus, the bioaccumulation xi i

lab1e 2. Recommended hioaccurnul ati an .factor re1 ations far stronti urn

Tax o H / Recommended Functional TiSSUF Environment Bisaccurnwlation Group Factor Re7 ati on

Fishes Bone ,411 waters Figure 1 F1 esh All waters Figure 1

A?gae Whol e A1 1 wa tiers zX1o3

Vascular plants (aquatic and emerg en t ) Whol e A1 1 waters 2x1o2 4 a Mol 1 uscs She1 1 A1 1 waters 6.8~10 /[ca], Soft tissues All waters 3x1 02 __ a .- /CalbV I: stable calcium concentration of water in ppm. xiii

0 R N L- WG 75- 4 44 7R 5 D

2 Id

5 0.1 0.2 0.5 1 2 5 10 20 50 100 CALCIUM CONCENTRATION IN WATER (pprn) for in Figure 7 Bioaccumulation factors strontium freshwater fishes as a function of calcium concentration in water. xiv

Tab1 e 3. Recommended bioaccumulatisn factor for tritium

~ _...__ . Taxon/ Recommended Functional Tissue Environment Bioaccumul ation Group Factor _I._s-I_.._....I._.. _. - ...... - A1 1 organisms All tissues A1 1 waters 1 xv

Table 4. Recommended bioaccumulation factors for stable iodine, iodine-129, and iodine-131 - - Taxon/ Recommended Functional Tisue Environment Bioaccumulation Group Factor Re1 ation

Stable Iodine and Todine-129 Fishes Muscle A1 1 waters 50 Ovary (eggs) 800 Thyroid ti ssue 290,000 C rus tacea A1 1 waters 340 Phytoplankton A1 ? waters 800

-..~Iodi ne-131

Fishes Muscl e A? 1 waters 40 Ovary (eggs) gooa Thyroid tissue 1 4 0,000 Aquatic insect 1 arvae Who? e AI 1 waters 400 1 USCS Soft tissues A1 1 waters 50 Mol She1 I 400 A1 gae Who1 e A1 1 waters 260 Macrophytes Who1 e A1 1 waters 120 -______aDefault value based on bioaccumulation factor of stable iodine. xv i factor of manganese in fishes is inversely proportional to manganese concentration in water [Equation (Z)]. Recommended bioaccumulati on factors are given in Table 5.

Cobal t

Cobalt in solution has a strong tendency to form complexes with dissolved organic matter. Since metals complexed with organic molecules have significantly lower availabilities to plants and animals than free ions, cobalt bioaccumulation factors would be expected to decrease with increasing eutrophy of water. Cobalt bioaccumulation factors, which in this report are based on cobalt concentrations in the soluble phase of water, conform to this expected pattern. Absorption efficiency of cobalt from food is very low in fishes.

This may explain the relatively low bioaccumulation factors of cobalt in fishes. Weeommended bioaceumulation factors for cobalt are given in Table 6. xvi i

Table 5. Recommended bioaccumulation factor relations for manganese

--- _I-._-._ ~~ Taxon/ Recommended Functional Tissue Environment Bioaccumulation Group Factor Relation Fishes Whol e A1 1 waters 6.7/[Mn] a Flesh A1 1 waters 0. 32/[MnYw

A1 gae Whol e A1 1 waters 5u bmer ged macrophytes Whol e A1 1 waters

F1oati ng-1 eaf macrophytes Whol e A1 1 waters 1Q3 Emergent vascular plants Who1 e All waters 7 o3

Mol 1 USCS : Bivalves Soft A1 1 waters c Shel 1

Snai 1s Soft Shel 1 Whol e

Crustaceans Who1 e All waters Insect larvae Whol e A1 1 waters

a lMnIb, = stable tnanganese concentration in water .in ppm. xvi i i

Table 6. Recamended bioaccurnulation factors for cobalt (based on isotope concentrations in water after filtration to remove particulates)

Taxon/ Recommended Functional Tissue EnLi ronment Bioaccuniul ation Group Factors

Fishes esha Mesotrophic and oligo- 32 0 F1 trophic waters Whol ea 448 Flesh Eutrophic waters 27 Who1 e 44 A1 gae Who1 e A1 1 waters 1 o4 Ernergm t vascular pl a'its Who1 e A71 waters I 03 Subrnerged- 1e?.f macrophytes Whol e MesotPophi c and 01 i go-" 1o4 trophic waters Eutrophic waters 2x1 o3 Floating-leaf Who1 e Mesotrophic and oliga- 1O3 macrophytes trophic waters Eutrophic waters 4x1 O2

1901 1uscs : BiVal ves Soft Mesotrophic and oliga- 1 trophic waters o4 Eutrophic waters 4x1 0' Snai, Soft All 7 5 waters 1o4 Insect 1arvae Mho1 e A1 1 waters 1o4 'lubi f i cid worms Whol e Eutrophic waters 5x1 O2 ce Whol e A1 waters C rust a an s 1 '1 O3

%Bioaccumulation factars of other tissues of fishes may be estimated from relative concentrations of cobalt given in Table 3.6.2. xix

CONTENTS Page 1 . Introduction ...... 1 1.1 Scope ...... 1 1.2 Objectives ...... 1 7.3 Format ...... 2 2 . Bioaccumulation Factor Concepts ...... 5 2.1 Idealized 6ioaccumulation Factor Patterns ...... 5 2.2 Physicochemical Forms of Raclionucl ides and Avai 1. ability to Biota ...... 9 2.3 Bioaccumulation Factors and Availability of Radio- nuclides ...... 10 2.4 Bioaccumulation Factors for Short-Lived Radio- nuclides ...... 11 2.5 Experimental Error in Determination of Bioaccumulation Factors ...... 14 2.5.1 Analytical Error ...... 74 2.5.2 Use of literature Values for Water Concentrations ...... 15 2.5.3 Filtration of Water Samples ...... 16 2.5.4 Determination of Organism Concentration Based on Dry Weight ...... 17 2.5.5 Averaging Isotope Concentrations of Different Tissues ...... 17 2.5.6 Determinations Made Under Nonsteady-State Conditions ...... 18 2.6 Use of Bioaccumulation Factoirs in Radiation Dose Estimation ...... 19 2.6.1 Dose to Biota ...... 19 2.6.2 Dose to Man ...... 20 3 . Bioaccmulation Factors for Cesium. Strontium. Tritium. Iodine. Manganese and Cobalt ...... 25 3.1 Cesium ...... 25 3.1.1 Cesium Metabolism ...... 25 3.1.2 Environmental Cesium ...... 28 3.1.3 Review of Cesium Bioaccumulation Factors ...... 32 3.1.3.1 Cesium Bioaccurnulation Factors for Fishes ...... 32 3.1.3.2 Cesium Bioaccumulation Factors for Algae ...... 44 xx

CONTENTS (continued) Page 3.1.3.3 Cesium Rioaccumulation Factors for Vascular Aquatic Plants ...... 48 3.1.3.4 Cesium Bioaccumulation Factors for Emergent Vascular Plants ..... 50 3.1.3.5 Cesium Bioaccuniwlation Factors for Invertebrates ...... 50 3.1.3.6 Cesium Bioaccumulation Factors for Amphibians ...... 50 3.1.3.7 Cesium Bioaccumulation Factors for Waterfowl ...... 50 3.2 Strontium ...... 63 3.2.1 Strontium Metabolism ...... 63 3.2.2 Environmental Strontium ...... 64 3.2.3 Review of Strontium Bioaccumulation Factors 64 3.2.3.1 Strontium Biaaccmulation Factors for Fishes ...... 66 3.2 .. 3.2 Strontium Bi oaccurnul ation Factors for Algae ...... 67 3.2,3.3 Strontium Biaaccumulation Factors for Aquatic and Emergent Macrophytes 78 3.2.3.4 Strontium Rioaccumulation Factors for Molluscs ...... 78 3.3 Tritium ...... 87 3.3.1 Hydrogen Bioaccumulation ...... 87 3.3.2 Pathways of Hydrogen Entry into Aquatic Organisms ...... $8 3.3.3 Compartment Size and Turnover Rates ..... 90 3.3.4 Tritium Concentration in Aquatic Food Chains ...... 93 3.3.4.1 Tissue-Water Tritium ...... 93 3.3.4.2 Tissue-Bound Tritium ...... 93 3.3.4.2.1 Exchangeable Ti ssue-Bound Tritium ...... 94 3.3.4.2.2 Nonexchangeable Tissue- Bound Tritfm ...... 96 3.4 Iodine ...... 103 3.4.1 Environmental Iodine ...... 163 xxi

CONTENTS (continued 1 Page 3.4.2 Iodine Metabolism...... - ...... - . . 103 3.4.3 Review of Iodine Bioaccumulation Factors ...... 105 3.4.3.1 Stable Iodine Bioaccumulation Factors for Fish Tissue ...... I 105 3.4.3.1 .I Stable Iodine Bioaccumulation Factors for Fish Muscle . . . . 109 3.4.3.1.2 Stable Iodine Bioaccumulation Factors for Fish Thyroids and Ovaries ...... 110 3.4.3.2 Lodine-131 Bioaccumulation factors for Fish Tissue . . . . e ...... 114 3.4.3.2.1 Iodine-131 Bioaccumulation Factors for Fish Muscle . . . . ?14 3.4.3.2.2 Iodine-131 Bioaccumulation Factors for Fish Thyroids . . 117 3 4.3.3 Stab1 e Iodine Bioaccumul ation Factors for Plants ...... - ...... 118 3.4.3.4 Iodine-131 Bioaccumul at i on Factors for Plants ...... I . . . 118 3.4.3.5 Stable Iodine Bioaccumulation Factors for Invertebrates . e ...... I . . 118 3.4.3 6 Iod i ne-1 31 B i oaccumul at i on Factors for Invertebrates . . . . e . . . . a . . 121 3.5 Manganese ...... 129 3.5.1 Environmental Manganese ...... 129 3.5.2 Manganese Metabolism ...... 130 3.5.3 Review of Manganese Bioaccumulation Factors a . . . 131 3.5.3.1 Manganese Bioaccumulation Factors for Fishes ...... - . . . 131 3.5.3.2 Manganese Bioaccumulation Factors for Plants ...... 136 3.5.3.3 Manganese Bioaccumulation Factors for Invertebrates ...... 140 3.5.3.3.1 Manganese Bioaccmulation Factors for Molluscs ...... 140 3.5.3.3.2 Manganese Bioaccumulation Factors for Invertebrates Other than Molluscs ...... 145 xxi i

CONTENTS (continued) Page 3.5 Cobalt ...... 151 3.6.1 Cobalt Metabol ism ...... 151 3.6.2 Environmental Cobalt and Availability ...... 153 3.6.3 Review of Cobalt Bioaccumulation Factors .... 154 3.6.3.1 Cobalt Bioaccumulation Factors for Fishes ...... 155 3.6.3.2 Cobalt Bioaccumulation Factors for Plants ...... 166 3.6.3.3 Cobalt Bioaccumulation Factors for Invertebrates ...... 169 Bibliography ...... 179 Appendix ...... 211 TP, B L ES Tab1 e Pap- 3.1.1 Biological half-lives and elimi ion coefficients for cesium and [k/(k+A)]'s for p4gCs ...... 26 3.1.2 Sorption of 134Cs to various concentrations of clay suspended in distilled water . . . . I . . . , e . . . * . . 29 3.1.3 Percentage of cesium in water in the particulate phase in some freshwater ecosystems ...... e a . . 30 3.1.4 Mean 137Cs bioaccumulation factors for whole fishes from Finnish lakes ...... I . . 34 3.1.5 Bioaccumulation factors for cesium as related to potassium and suspended solids concentrations ...... 36 3.1.6 Bioaccumulation factors far cesium in fishes ...... 37 3.1.7 Bioaccumulation factors for cesium in freshwater algae . . . 45 3.1.8 Bioaccurnulation factors for cesium in aquatic vascular plants...... , . . . '...... 49 3.1.9 Bioaccumulation factors for cesium in emergent vascular plants., ...... , ...... 51 3.1.10 Bioaccumulation factors for cesium in aquatic invertebrates.. . . . ".. . *. . . (...... 52 Bioaccumulation factors for in amphibians 3.7.11 137Cs . . . a . . 54 3.1.12 Bioaccumulation factors for 137Cs in waterfowl ...... 55 3.2.1 Percentage of radiostrontium in water in the particulate phase in some freshwater ecosystems ...... , . . . 65 3.2.2 Bioaccumulation factors and discrimination coefficients for strontium in fishes ...... 68 3.2.3 Parameters from linear regressions between BF(Sr)i and rCalw for bone and flesh of fishes ...... I . . 75 3.2.4 Bioaccumulation factors and discrimination coefficients for strontium in algae ...... - 76 3.2.5 Bioaccumulation factors and discrimination coefficients for strontium in aquatic and emergent macrophytes . . . . 79 xxiv

TABLES (continued)

----Table Page 3.2,s 13ioaccumulatjon factors and discrimination coefficients for strontium in mollusc shells ...... 80 3.3.1 Component s-ime and so~lrcecontri butl’or; for a hypothetical fish...... 92 Steady-state in 6rgan-i~groups 3-3.2 concentration cornon in living tissue ...... 95 3.4.1 Iod:’ne concentrations in surface waters the ocean, and geologic SOU~CPS...... 104 3,4*2 Elimination rates for iodine-I31 in aquatic vertebrates. .. 106 3*4*J Stable iodine in freshwater fish muscle ...... 137 3.4.4 Stable iodine in anadromous and marine fish muscle ..... 1os Total present the thyroid eggs of five 3-4.5 iodine insteelhead troutand Rainbow and California ...... 112 3-4.6 Thymid stable iodine bioacc.mulation factors for m2rine fish ...... 113 3”4.7 Stable iodine bioaccumulatfon factor for fish ovaries (eggs) derived frcm relat-ive concentratfons of stable isdine in muscle and ovar-ies ...... 115 31 biaaccmulation factors far muscle fresh- 3,4.8 Iodine-l in watw animals ...... 116 3-4.4 Iadine-131 bioaccmulation factors for freshwater plants ...... 119 Stable iodine factors for 3.4.10 bisaccumulation invertebrates . . 120 Iodine-131 factors 3-4-11 biaaccumulation for invertebrates .... 122 Bioaccuinulat’ion for freshwater 3.5.7 factors manganese in fishes ...... 132 Manganese concentrations var-s’ata’rdn 3.5.2 and coefficients of in wate~and freshwater fishes ...... 135 3.5.3 Bisaccumulatjon factors Poi- ~anganesein aquatic plants . . 138 xxv

TABLES (continued) Table Page

L 3.5.4 Bioaccumulation factors for manganese in molluscs ..... 141 3.5.5 B i oaccumu 1a ti on factors for manganese in invertebrates other than molluscs ...... 146 3.6.1 uiological elimination rates for 60~oin various aquatic organisms ...... 152 3.6.2 Relative steady-state concentrations of 6oCo in ti sues of Ictalurus melas (black bullheads) from White Oak Lake ...... 156 3.6.3 Bioaccumulation factors for cobalt in fishes from eutrophic environments (based on filtered water concentrations) ...... 157 3.6.4 Bioaccumulation factors for cobalt in fishes from mesotrophic environments (based on fi 1 tered water concentrations) ...... 161 3.6.5 Mean bioaccumulation factors for cobalt in fishes from mesotrophic and eutrophic environments ...... 163 3.6.6 Bioaccumulation factors for cobalt in aquatic plants .... 167 3.6.7 Mean cobalt bioaccumulation factors for submerged and floating-leaf vascular plants from Perch Lake and Lake Maggiore (based on filtered water concentrations) .) . . 170 3.6.8 Bioaccumulation factors for cobalt in invertebrates (based on filtered water concentrations) ...... 171 FIGURES Figure Page 3.1 .l Bioaccumulation factors for 137~sand stable cesium in freshwater fi'shes as a function of potassium concentration in water ...... 43 3.2.1 Bioaccumul ation factors for strontium in freshwater Pi shes as a function of calcium concentration in water ...... 74 3.3.1 Pathways of hydrogen entry into the hydrogen components of ~RIaquatic animal ...... 89 Three-link food diagram with organisms 3.3*2 chain compartmentalized ...... 97 3.5.1 Bioaccumulation factors for manganese in freshwater fishes as a function of manganese concentration in water ...... 137 3.6.1 Concentration of stable cobalt in fish flesh in various environments ...... 165 1. INTRODUCTION

1.1 Scope

This report analyzes over 200 carefully selected papers to provide concise data sets and methodology for estimation of bioaccumulation factors for selected radionuclides in major biotic components of freshwater environments. The papers were critically reviewed from the standpoint of experimental techniques for each of the selected radionuclides. Since conditions in the laboratory are often unnatural in terms of chemical and ecological relationships, this review is restricted as far as possible to bioaccumulation factors determined for natural systems. Cesium, strontium, iodine, manganese, and cobalt were chosen based on an ORNL survey of releases from 16 light-water-coaled nuclear power stations. This survey showed that the major radiological impact on the aquatic biota and man comes from these radionuclides (Kaye, 1973). Although tritium contributes only a minor fraction of the dose from liquid effluents, it is nonetheless treated too because of a concern that it may be concentrated in passage through food chains. Bioaccumu- lation factors for freshwater environments are of immediate concern since most nuclear plants (operable, under construction, or planned) are located adjacent to bodies of freshwater.

1.2 Objectives

The objective of this report is to present bioaccumulation factors for the selected radionuclides in the different tissues of major biotic components of freshwater systems. The bioaccumulation factor of a radionuclide in an organism or tissue ma.1 exhibit wide variations among sites 2 that are related to differences in stable-element and carrier-element concentrations and other limnological variables. For this reason, bioaccumulation factor relations rather than single bioaccumulation

factors are presented to account for site-specific variations ~

1.3 Format

General discussions of important aspects of element bioaccumulation are presented in Section 2, We present simple models to illustrate the relation between bioaccumulation factors for radionuclides and the concentrations of their stable-element and carrier-element analogues. The effects of physicochemical form and other factors in causing deviations from these models are also discussed along with methodology for converting stab1e-element bioaccumulation factors to bioaccumulation factors for- shorter-! ived rada’onuclides. Section 3 reviews the 1 i tera- ture on bioaccumulation factors for the six selected radionuclides and gives recommended bioaccumulation factor relations based on the literature values. Reference

Kaye, S. V. , Assessing potential radiological to aquatic biota in response to the National Ehvironmental impactsPolicy Act (NEPA) of 1969. IN Symposium on Environmental Behavior of Radionuclides Released in the Nuclear Industry, IAEA Symposium, Aix-en-Provence, France, 14-18 May 1973.

2. BIOACCUMULATION FACTOR CONCEPTS

The bioaccurnulation factor of an organism or tissue i is the steady-state ratio of radionuclide concentration in the organism or tissue to that in water:

where

EF(R)i = bioaccumulation factor for radionuclide R in organism or tissue i, [Rli = radionuclide concentration (gCi/g fresh weight) in organism or tissue i, and [R], = radionuclide concentration in water (pCi/g), a constant.

Bioaccumulation factors are used to predict radionuclide concentrations in whole organisms or their tissues from knowledge of the radionuclide concentration in water for chronic releases of radionuclides. Bioaccu- mulation factors also appear in expressions that are used to predict the transient dynamics of radionuclide concentration in organisms

(Vanderploeg et a7 e 1974). Therefore, an understanding of variables that affect bioaccumulation factors is central to understanding steady- state and transient dynamics of radionuclide concentration in aquatic organisms.

2.1 Idealized Bioaccumulation Factor Patterns

Bioaccumulation factors for ra'3ionuclides are often portrayed as being related to ambient concentrations of their stable or non-isotopic 6 carrier element analogues according to three idealized patterns (e.g., Fleishman, 1973; Peterson, 1970). The first pattern is that the bioaccumulation factor for radionuclide R in organism or tissue i, BF(R)i, -t is constant regardless of stable-element or carrier-element concentrations:

BJ-(I?). = const. (2.1 .l) 1

The second pattern is that the concentration of an element in an organism or tissue is homeostatically maintained at a constant concentration despite different concentrations of the element in water. This implies that the biaaccurnulation factor of a homeostatically control led stable element is inversely proportioned to its concentration in the water:

where Ei = concentration of stable element .in organisms or tissue i, a constant (pg/g fresh weight), and

= concentration of stable eleiiient in water (pg/g).

Assuming that the specific activity, that is, the ratio of radionuclide concentration to the stable element concentration (pCi/g stable element) in i is equal to that of the water, the bioaccumulation Factor for the radionuclide, I3F(RIi, is also inversely proportional to the concentration of the stable element:

BF(R)i = Ci/cCl, . (2.1.3)

Taking the natural logarithm of each side of Equation (2.1.2) and (2,7.3) results in the linear relation: 7

In BF(R)i = In BF(C)i = ln ci - ln[C], . (2.1.4)

Equation (2.1.4) implies that the plot of ln 5F(RIi \~ersusln[C]w has a slope of -1. The third pattern is that the bioaccurnulation factor for the radionuclide, BF(R) i, is inversely proportional to the concentration of a non-isotopic carrier element in water, a non-isotopic carrier element being nearly chemically similar to but occurring in higher concentrations than the stable-element analogue The derivation of this relationship follows. The bioaccumulation factor for radionuclide R is related to BF(C*Ji, the bioaccurnulation factor for the carrier element C*, by

(2.1.5) which may be written as

(2.1.6)

where (R/C*) and ( R/C*)w are ratios of radionuclide concentration to carrier element concentration found in the organism or tissue and in the water, respectively. Assume that the non-isotopic carrier element is horneostatically controlled in the organism, that is: 8 where

Zi* = concentration of non-isotopic carrier element in organism or

tissue i, a constant (pg/g fresh weight), and [C*], = concentration of non-isotopic carrier element in water (pglg).

Combining Equations (2.1.6) and (2-1.7) and letting qi = (R/C*)i/(R/C*)w,

(2,l .8)

Taking the natural logarithm of each side of Equation (2.1 .8),

Since Xi* is a constant, a plot of ln BF(R)i versus ln[C*], will have a slope of -1 if qi is constant.

If water were the immediate source of the radionuclide and carrier element to organism or tissue i, qi would be called the "observed ratio" (Comar et a1 . , 1956). Since water may not necessarily be the direct source of an element to an organisill, we will designate qi the discrimination coeff ici ent . On the basis of empirical data on 137~sand its carrier element potassium, Fleishman (1973) suggested that qi is not a constant but rather a function of [C*],. It can be shown niathematically that if an animal accumulates the carrier and nuclide from water and food, qi beconies a function of [C*], (see Appendix).

Because the number of sites where steady-state bioaccuniul ation factors can be determined from the release of radionuclides is limited, bioaccurnulation factors for the radionuclides are often estimated from distributions of their stable element analogues. Therefore, the requirement of equal specific activities in organisms and water that we have for Equation (2.1.3) applies as well to Equations (2.1.1) and

(2.1.8) or any relation when stable element data are used to estimate bi oaccumul ation factors.

2.2 Physicochemical Forms of Radionuclides and Availability to Biota

Radionuclides exist in a wide variety of physicochemical forms in natural waters and their different physicochemical forms have different availabilities to aquatic biota. The major physicochemical forms of trace metals that Gibbs (1973) reports for rivers illustrate this diversity:

1. Dissolved ionic species and inorganic associations, 2. Complexes with organic molecules in solution, 3. Adsorbed to solids, 4. Precipitates and coprecipitates on solids (metallic coatings), 5. Incorporated in solid biological materials, and

6. Incorporated in crystalline structures.

Both unicellular and multicellular algae, which generally form the base of the aquatic food web, accumulate elements from the soluble phase. Likewise, some radionuclides are directly accumulated by aquatic animals from the soluble phase. Aquatic vascular plants, which may be important as food to consumers in some systems, accumulate elements from the soluble phase of water and also frori the interstitial water of sediment. It has also been demonstrated that radionuclides and metals complexed 10 with organic molecules may have significantly lower availabilities ta algae and animals than free ions (Timofeeva et al., 1960). Aquatic animals, especially filter feeders, may accumulate radionuclides from the suspended phase. The availability sf different forms of radionuclides in the suspended phase has not been studied. It can be said that adsorbed (exchangeable) radionuclides and radionuclides in solid biological materials are available to some aquatic consumers.

Elements in the crystal structure of clay minerals and other lithogenous detritus are almost completely unavailable.

Radionuclides enter food webs not only from the water but also from bottom sediments, and availability of radionuclides to the food web varies with sediment type. Benthic invertebrates accumulate radionuclides from bottom sediments. Fishes may accumulate radionuclides indirectly from bottom sediment by ingestion of these invertebrates and also directly by incidental ingestion of sediment with prey

(Gallegos et al , , 1970). Absorption of radionuclides from ingested sediment varies with the nature of the sediment (Gal7Fzgos et al., 1970; Eyan and Kitchings, 1974). In addition, different sediments have different capacities for sorption of radisnucl i des (Friend,

1963 ; Imieni ck and Gardi ner 1965) .

2.3 Bioaccumulation Factors and Availability of Radionucl ides

For those elements that are horneostatical y rnainta ned at constant concentrations in a given organism, the concentration of stable element in the organism is independent of concentrations of stable element in water or its availability fram prey, sediment, and different physicochemical forms of the water. If the radionuclide in the organism has 11

the same specific activity as that in waterS the bioaccumulation factor for the radionuclide will be given by Equation (2.1.3). In contrast, differences in availability of radionuclides in different sediments and different physicochemical forms in water can lead to marked deviations from the idealized pattern of a constant bioaccumulation factor, Equation (2.1 .I). For the bioaccumulation facto,@ pattern of Equation (2.1.8), we pointed out that qi is a function of [C*Iw. We note too that if sediment is an important source of radionuclide to the animal, the concentration of carrier in sediment relative to food and water will have an influence on qi (see Appendix). Moreover, variations in the proportions of radionuclide and carrier element appearing in to a, various physicochemical forms can lead variations in i sr’nce the radionuclide or carrier could be preferentially accumulated by the food web.

2.4 Bioaccumulation Factors for Short-Lived Radionuclides

Because of radioactive decay, the bioaccumulation factor for a short-lived radionuclide is expectEd to be less than the corresponding bioaccumulation factor of the stable element. Since the bioaccumulation factors of some stable elements have been more corrmonly measured than the bioaccumulation factors of short-lived radionuclides, a methodology is presented to predict the bioaccurnulation factor for the short-lived radionuclides from the bioaccumulation factor of the stable element, BF(C)?. The methodology requires knowledge of the elimination coefficient, k, in organism or tissue i. 12

Treating organism or tissue i as a single compartment, the rate of change in the amount of radionuclide R in i, Ri, ($i)is

dRi -_ - IT - Ri (k +- A), (2.4.1) dt and the steady-state amount of stable element C is given by

, I, 1 c. = (2.4.2) 1 k where

IT = rate of input of radionuclide from all sources (pCi/day),

IT' = rate of input of the corresponding stable element from all sources (yg/day),

k = elimination coefficient at steady-state concentration of C in i (day-'), and

h = radioactive decay constant (day-').

Assuming constant weight. of the organism or tissue, the steady- state concentration of radionuclide R and stable element C are given by

I (2.4.3) =

(2 4.4) where B = biomass or fresh weight of the organism or tissue i (9). Assuming all uptake is from water,

(2 "4.5) 13

(2.4.6)

where a = uptake coefficient for element C from water (day-’ ) ,

[RIw = radionuclide concentratior in water ($i/g) and

[C], = stable element concentration in water (p.g/g).

Dividing Equation (2.4.5) by [Rlw ard Equation (2.4.6) by [C], gives

(2.4.7)

a BF(C) . = ~j (2.4-8) 1 which implies that

for uptake from water. The derivatim of Equation (2.4.9) parallels that given by Peterson (1970). Thus if the radioactive decay constant,

A, is significant relative to the elimination coefficient, k, BF(R)i will be significantly less than BF(C)i. If there are intervening steps

in the food chain between the organism or tissue and water, radioactive

decay will occur at each step and further diminish BF(R)i. Since the e1 imination coefficients of prey (or forage) are generally much larger than the elimination coefficients of their respective predators, Equation

(2.4.9) will generally not greatly overpredict 5F(R)i. Since Equation

(2.4.9) will not greatly overpredict the BF(R)i and since information on

food web structure is needed to derive a more accurate BF(R)i, we suggest that Equation (2.4.9) be used to convert bioaccumulation factors for stable elements or long-lived nuclides to bioaccumulation factors for 14

k short-lived radionuclides. Tables of k and [m]values are given when the bioaccumulation factors for shorter-lived isotopes must be estimated from the bioaccumulation factors for stable elements.

2,5 Experiniental Error in Determination of Bioaccumulation Factors

In earlier sections we treated the effects of variations in certain environmental conditions on bioaccumulation factors. Varia- tions a1 so resul t from experimental error , that is , errors resul ting from artifacts of analysis, eval uation and presentation of data, Below we discuss some important sources of experimental error and the precautions which can be taken to minimize them. These important sources of experimental error are: a. Analytical error , b, Use of literature values for water concentrations, c. Filtration of water samples, %, Determination of organism concentration based on dry weight, e. Averaging isotopic concentrations of different tissues, and f. Determinations made under nonsteady-state conditions.

2.5 1 Analytical Error

The concentrations of radionuclides or stable elements in most enV ronmental samples are generally so low that serious problems can be encountered in their measurement, especially in water. Eight laboratories participated in an intercalibration study involving spiked freshwater samples (Maletskos, 1972). The results were given as a ratio of the reported values to the ”correct” values 9- the 15

standard dev ation. The resu ts were: 54Mn = 0.91 -f 0.20, 6oCo = 0.97 t 0.40, 137Cs = 1.05 -f 0 27,, and 'OS, = 1.09 -I- 0.46. Iodine and tritium were not included in this study. Although it is not likely that errors n analysis of b ological material would parallel those occurring in water analysis, it is obvious that analytical error can result in a significant variation in measured bioaccumulation factors. Significant error may also be associated with measuring stable element concentrations in water, especially for cesium, cobalt and manganese. Unfortunately, it is not generally possible to evaluate the accuracy of measurements from particular studies. In deriving our bioaccumulation factor relations, we imp1 ici tly assumed that variations in bioaccumulation factors arising from analytical error were small in relation to variations caused by environmental variables and that deviations caused by analytical error were randomly distributed.

2.5.2 Use of Literature Values for Water Concentrations

Some of the variation in reported bioaccumulation factors can be attributed to the use of average literature values for stable isotope concentration in water. This is especially true in freshwater environments where element concentrations are more variable than in marine environments. Thompson et a1 . (1972) derived some average bioaccumulation factors by determining the average stable element concentration in organisms from a number of freshwater habitats and then dividing this average concentration by the average stable element concentration of a number of freshwater habitats. Often the water values were taken from studies different from those giving the concentrations in the .P 16 organisms, Thus, some error may have been introduced into the average bioaccumulation factor values. In selecting or calculating bioaccumulation factors for this report only those studies that collected water and organism samples from the same locality are included.

2.5.3 Filtration of Water Samples

Some experimenters have reported bioaccumulation factors based on concentrations of isotopes from filtered as we1 1 as unfiltered water samples. Since a significant fraction of some elements in water may be in the suspended phase - in some cases greater than 905 - hioaccumulation factors based on concentrations in fil ‘cered samples may be much 1 arger than bioaccumulation factors based on unfiltered concentratians.

So understand the significance of filtered or unfiltered bioaccumulation factors for prediction at proposed nuclear power plants it is necessary to discuss the nature of the effluent and how radionuclide concentrations in the water at the proposed reactor sites are estimated.

Radionuclides in the reactor effluent are in the soluble phase. Concen- trations in the body of water of concern are usually estimated from predicted concentrations in the effluent and from water turnover rates.

Actual concentrations in the water may be somewhat lowet- because the bottom sediment may serve as a sink far radionuclides. Since radionuclides may become distributed between soluble and suspended phases, the “predicted concentration” approximates the unfiltered concentration of the radionuclide in the water. Assuming that the specific activity 17 of the radionuclide in the soluble and the suspended phases are equal, an unfiltered bioaccumulation factor will most closely predict the radionuclide concentration in the organism. Generally, filtered bioaccumulation factors will overpredict the radionuclide concentration in organisms.

2.5 -4 Determination of Organi sm Concentration Based on Dry Weight

Bioaccumulation factors expressed in terms of dry weight of the organism are higher than those calculated using wet weight concentrations. Thus, care is taken to ensure that bioaccumulation factors listed in this report are on a fresh weight basis. When dry weight bioaccumulation factors were given in a publication, they were con- verted to fresh weight bioaccumulation factors if conversion factors between fresh weight and dry weight were available in the same publication. In some cases where conversion factors were not available in the same publication, conversion factors from other sources were used

2.5.5 Averaging Isotope Concentrations of Different Tissues

In reporting or evaluating bioaccurnulation factors care must be taken to specify what tissue or tissues of the organism were analyzed. Certain elements have an affinity for different tissues. For example, strontium concentrates to a higher degree in bony tissues, and iodine concentrates to a higher degree in the thyroid than in other tissues of the organism. Thus, any bioaccumulation factor is incomplete without reference to the specific tissue analyzed. The whole-body bioaccumulation factors, which are often reported, may be different from those of any particular tissue" Another potential source of uncertainty associa,ted with whole-body bioaccumulation factors is that the organism's gut may contain sediment when the whole-body bioaccumulation factor is determined. The concentration of metals, such as cesium, manganese, cobalt, and strontiuiii in sediment may be significantly higher than the concentrations found in the tissues of the organism. Sediment contamination is most likely a significant factor for benthic invertebrates.

2 5.6 Detemi nations Made Under Nons teady-State Condi t on s

Measureiiients of organism concentrations in situations where a steady-state condition is not present can result in low or high bioaccumulation factors depending on the circumstance. If an organism is transferred from an uncontaminated environment to an environment with a fixed level of a radioisotope, the time necessary for the establishment of a nearly steady-state concentration is a function of the biological half-time of that element in the organism and the radioactive ha1 f-life. Obviously, organism concentrations analyzed prior to the establishment of a steady state will result in low bioaccumulation factors. If the environmental concentration is subjected to sudden variations, the subsequent changes in the organism concentrations may lag behind that of the environment. Low bioaccumulation factors would be expected if determinations are made shortly after a sudden rise in environmental concentration, or after an organism mves from an area of law isotope concentration to an area of high concentration. The opposite situations would result in high bioaccumulation factors. 19

2.6 Use of Bioaccumulation Factors in Radiation Dose Estimation

The two principal uses of bioaccumulation factors for aquatic biota are: (1) calculation of radiation dose to biota from radionuclides deposited in tissues, and (2) calculation of radiation dose to man from consumption of contamiqated biota. In both situations the radiation exposure is within the body and not exterior to it.

2.6.1 Dose to Biota

The internal absorbed dose (r'adslyear) to organism i from an internally deposited radionuclide is estimated from the energy deposited per gram of tissue and can be expressed mathematically as:

Di = [R] BF(R)i E K (rads/year) (2.6.7) W where

Di = internal absorbed dose (rads/year) to organism i from a radionuclide uniformly deposited in its tissues,

[R], = radionuclide concenYation in water (pCi/g), BF(R)i = bioaccumulation factor for radionuclide R in organism i,

E = effective absorbed energy (MeV) of the radionuclide for the physical dimensjons appropriate for the organism i, and

K = conversion factor to convert $i/g to radsjyear.

There is no radioactive decay correction in the above equation because the concentration of the radionuc1"de in water is assumed to remain constant and the concentration in zhe biota is in a steady-state with 20 the water. In actual practice, the average annual concentration of the radionuclide in the receiving water at the paint of interest is used in dose calculations.

The variable K in the above equation is defined as a conversion factor, but it can represent any type of internal dosimetry model.

The variables BF(R)i, and E are always required, no matter how specific or general the conversian factor K.

2.6.2 Dose to Man

A bioaccumulation factor is used to estimate the concentration of a radionuclide in the aquatic food (organism or tissue i) consumed by man. The intake rate of a radionuclide ($i/day) by man is calculated from

Ii = [RIbV BF(RIi Gi (pCi/day) (2.6.29

Wklb re

Ii = man's daily dietary intake rate of a radionuclide (pCi/day) contained in aquatic food organism or tissue

[Rlw = radionuclide concentration in water (pCi/g) ,

BF( RIi = bioaccumulation factor for radionuclide R in organism or tissue i , and

Gi = man's daily intake rate of organism or tissue i (g/day).

In most exposure situations each organisiii is contaminated by more than one radionuclide, and inan's dietary intake usually includes more 21 than one type of aquatic biota. Therefore, a summation must be per- formed over all radionuclides and biota to estimate man's total intake of radioactivity. The internal dose to a reference organ or to the total body of man is calculated using the estimated radionuclide intake rate (pCi/ day) for each radionuclide in addition to the following dosimetry factors: fraction of the intake deposited in the organ of reference, the effective elimination constant for each radionuclide in the organ of reference, the effective absorbed energy of each radionuclide in the organ of reference, and the mass of the reference organ. 22

References

Comar, C. L., R. H. Wasserrnan, and M. Nold, Strontium-calcium discrimination factors in the rat,M. Proc. Soc. Explt. Biol. Med. 92, 859-863 (1956).

Eyman, I_. D. and J. T. Kitchings, Availability of 137Csto fishes from ingested, IN Proceedings of the XIX Congress of the International Association of Limnology (SIL) , Winnipeg Canada, August 22-29 (1974). (in press)

Fleishman, D. J., Accumulation of artificial radionuclides in freshwater fish, IN Radioecology, V. M. Klechkovskii, G. G. Polikarpov, and R. M. Aleksakhin (eds.), pp. 347-369, John Wiley and Sons, New York, 1973.

Friend, A. G., The aqueous behavior of strontium-85, cesium-137, zinc-65, and cobal t-60 as determined by laboratory type studies, IN Transport of Radionuclides in Freshwater Systems, B. H. Kornegay, et a1 . (eds .) , pp. 43-60, TID-7664 (1963). Gallegos, A. F., F. W. Wicker, and T. E. Hakonson, Accumulation of radiocesium in rainbow trout via a non-foodchain pathway, IN Proceedings of Fifth Annual Health Physics Society Midyear Topical Symposium: Health Physics Aspects of Nuclear Siting, VO~ 2, pp. 477-497, AEC CONF-701106 (1970). Gibbs, R. J., Mechanisms of trace metal transport in rivers, Science 18Q, 71-73 (1973).

Lomenick, T. F., and D. A. Gardiner, The occurrence and retention of radionuclides in the sediments of White Oak Lake, Health Phys. -11 , 567-577 (1965). Maletskos, C. J. , Intercalibration and intercomparison study of radionucl ide concentrations in Lake Michigan samples, New England Deaconess Hospital, Progress Report October-March, 1972, COO-3371-1, USAEC Contract AT( 11 -1 ) -3371 . Peterson, H. T., Jr., Stable element and radionuclide reconcen- tration in aquatic ecosystems. I. Stable element effects in the equil ibrium situation, Fifth Annual Health Physics Society Midyear Topical Symposium, Idaho Falls, Idaho, November 3-6, 1970.

Thompson, S. E. , C. A. Burton, D. J. Quinn, and Y. C. Ng, Cancen- tration factors of chemical elements in edible aquatic organisms, Lawrence Livermore Laboratory, Univ. California, Livermore, 1972. 23

Timofeev - Resovskii, N. V., E. A. Timofeeva - Resowskaya, G. A. Milyutina, and A. B. Getsova, Coefficients of accumulation of radioisotopes of sixteen cifferent elements by freshwater organisms and the effect of EDTA on some of them, Dokl. Nauk SSSR -132, 369-372 (1960). Vanderploeg, H. A., R. S. Booth, and F. H. Clark, A specific activity and concentration model applied to cesium movement in an oligotrophic lake, IN Proceedings of Symposium on Mineral Cycling in Southeas tern Ecosys terns, A.ugusta , Georgi a, May 1-3, 1974 (in press).

ULATION FACTCRS FOR CESIUM, STRONTI TRITIUM, IODINE a MANGANESE AND COBALT

3.1 .I Cesium Metabolism

Because of their chemical similarities, cesium and potassium are metabolized somewhat similarly in organisms. In freshwater ec~systems, stable cesium, 133~s,is an element havjng concentrations ranging between

.Oil and 1.2 ppb for lakes and rivers (Capeland and A.yersy 15370; K~~~~i~i~~n~n~

1972; Kolehmainen and Nelson, 1969; Merlini et al., 1967; el Scan 7 967). tassiurn -is more abundant ‘in freshwater than cesium atid concentrations usually range between 0.2 ppm dnd 1 The bioaccumulatian fxtor for cesfu pendent af cesium concentrations found in natural bodies sf water (Gertr,

1973; Kfaag, 1964; Kolehrnainen et al. I 968; ~~~e~~ain~4~~1972; Williams 26

Table 3.1.1 Biological half-times and elimination coefficients for cesium and [k/(k+X)] values for 136Cs

Biological k Taxon/functional k - for half-time (days-l 1 kt- Reference category (days 1 136X CS Uni cell ul ar a1 gae 1 0.69 0.93 1.50 Mu1 ti cell u7 ar a1 yae 2 0.35 0.88 1.47 Aquatic vascular plants: f 1 oati ng 20 0.035 0.40 1.37 rooted 60a 0.012 0.18 1.37 Zooplankton 5 0.14 0.74 1.46 Benthi c insect 1 arvae 7 0.10 0.67 1.25 Cl ams 40 0.01 7 0.24 1.28 Fishes 100 0 8069 0.12 1.13, 1.23, 1.24, 1.45

%ince rooted plants may accumulate part of their burden of cesium from interstitial water of bottom sediments, time to steady-state may be niuch greater than if uptake were entirely from water above sediments. 27 biological half-times and eliminatim rate coefficients values for cesium in different taxa or functioqal groups. These coefficients must be considered as approximate since they may be influenced by such factors as temperature, size, growth rates, feeding rates, and species differences (Vanderploeg et a1 . , 1974). Most bioaccumulation factors, except those for algae, were derived entirely from stable

cesium or from 137Cs (30-year radioactive half-1 ife) concentratians ~ Cesium-134, having a 2.2-year half-life, will not have an appreciably different bioaccumulation factor from that of stable cesium or 137Cs. Cesium-136, which has a 13-day radioactive half-life, will have an appreciably lower bioaccumulation factor than those of the longer- lived isotopes in some biota; therefore, Table 3.1.1 includes a set of approximate [k/(k+A)] values to c.onvert bioaccumulation factors for stable cesium, or 137Cs, to bioaccumulation factors for 136Cs.

The primary mode of accumulation of cesium and potassium in fishes is generally thought to be via absorption from food (Kolehmainen, 1972); absorption Pram food may also be a primary mode of uptake for many aquatic invertebrates. However absorption of cesium From ingested sediments may be siZjnificant in some systems (6allegos et a1 ., 1970). In fishes (Kolehmainen, 1972) and other animals (Pendleton et al., 19651, the absorption efficiency of potassium and cesium from food varies but is often high--nearly 100 percent for soqe foods. The excretion coefficient k*ll of potassium is on the order of 3 to 4.5 times greater than the excretion coefficient, k, of cesium in fishes (Kolehmainen, 1972) and in terrestrial animals (Fujita et al. 1966, Pendleton et al. 1965), Pendleton pointed out that this will result tn a cesium to potassfurn ratio in the 28

predator that is higher thhan that of its prey by a factor of about three,

assuming that absorption efffciency of cesium is the same as that for potassium. Assuming that the potassium concentrat-ion in the predator is

the same as that in its prey, the bioaccumulation factor for cesium should

increase by abaut. a Factor of time with each trophic level. Note that

the variable k*/k appears in our general expression for q (see i Appendix).

3,1 .2 Enviroriinental Cesium In solution, cesi!,rm and potassium: like athw alkali metals, exist

primarily as free ions and do riot form inorganic complexes. Because cesium

-is st~snglyadsorbed by suspended particulate materials, especially clays

these materials may remove it from the soluble phase tu varying degrees.

Potassium, too, is sorbed, but its distribution coefficient, Kd, on SIJS-

pended sediment (that. is the concentration sf element sorbed tu sediment

per concentration in water) is iimy tines ssnai’ler than that of cesium

(Reynolds and G1 oyna 1364).

Table 3.1.2 shows that the percentage of ces’ium in the particulate

phase increases with clay partl’cle concentration, For freshwater ecosystems

the fraction of cesium in water in the suspended phase ranges between l9 and 92 percent, and this fraction appears to be roughly correlated with concentration solids of suspended (Table 3.1.3) kctiuse potassium is

less strongly sorbed thm cesium, the percentages OQ patassium appearing

in the suspended phase are very much snialler than those for cesium ?’VI these bodies water, of Since algae accumulate cesium, potassium, and other elements from the soluble phase, the availabilities to the Puod chain of cesium

and of cesium relative to potassium becoiiie. a function of suspended solids concentration and the distribution coefficients of cesiura? and potassium on 29

Table 3.1.2 Sorption of 134Cs to various concentrations of clay suspended in distilled water (Garder and Sku1 1berg 1964)

Concentration Percent of clay of cesium PPd sorbed Kda

16 20 15,600 32 31 13,600 64 43 11,800 128 53 10,800 256 65 7,300

aConcentration of element (cesium) sorbed to clay per concenrration in water. 30

Table 3.1.3 Percentage of cesium in water in the particulate phase in some freshwater ecosystems

Mean ‘percent in Suspended solids Separation kocati parti cul ate (ppm) or nature Reference on phase: of water body method White Oak Lake, 69 Turbid reservoir Ultracentrifu- 1.44 Tennessee gation: Par- ticles > (3.7 p Clinch River, 82 -92 25-1 85 U1 tracentri fil- 1.44 Tennesseea gatian : Par- ticles > 0.7 Tennessee River, 19-30 9-22 U1 tracentri fu- 1.44 Ten n es s ee a gati~n: Par- ticles > 0.7 1-1 Lake! Maggi ore, 28 Mesotrophi c Lake fi 1 ter 1.35 Italy 0.7-p b Experimental pond 58 21 0.45-p fi1 ter 1.10

a Range of means of three river stations.

bBased on five water values sampled between 24 and 80 days after introduction of radi onucl i des. 31 the solids. Thus, the qi values and bioaccumulation factors of cesium become functions of suspended-sol~ids concentrations and the distribution coefficients. Since the proportions of cesium and of cesium relative to potassium in the soluble phase are expected to vary greatly among environments, the bioaccumulation factors for cesium in algae and the qi values and bioaccumulation factors of animals may vary greatly among bodies of water having the same potassium concentrations. We shall see in the following sections that much of the scatter in bioaccumulation factors for cesium is apparently related to the proportion of cesium in the particulate phase.

The food web accumulates cesium not only from the soluble phase of water but also from suspended and bottom sediments. Filter feeders may accumu ate cesium adsorbed to particulate matter. Benthic invertebrates obtain cesium adsorbed to ingested bottom sediments. Fishes may accumulate ces i urn from bottom sediment by ingestion of these invertebrates and also by i nges t on of sediment with prey (Eyman and Kitchings, 1974; Gallegos et al.,

1970). Gallegos et al. (1970) reported that concentration of 137Cs in trout in a montane lake was higher than could be explained by the trouts' accumulation from their prey. Their experiments indicated that this difference could be explained by the trout ingesting small amounts of sediment. In fishes, absorption efficiency of cesium from ingested clay varies with clay

type: 8%for illite, 65% for kaolinite, and 85% for montmorillonite (Eyman and Kitchings, 1974). Since cesium is available from ingested sediments, differences in sediment type among water bodies may be another contributor to the variance in bioaccumulation factors among bodies of water. 32

3.1.3 Review of Cesium Bioaccumulation Factors This section presents bioaccurnulation factors and bioaccumulation factor relations for the fsl lowing categories of aquatic biota: 1. Fishes 2. Algae 3, Macrophytes (vascular aquatics) 4. Emergent vascular plants 5. Invertebrates-,-mQlluscs, insects, and crustaceans 6. Arnphibians 7. Water-fowl Except for algae, all bioaccumulation factors for cesiutii were derived from distributions of stable cesium or from steady-state distributions of

137Cs in natural or semi-natural bodies of water. Since ecological relationships in aquaria do not completely represent the natural system, bio- wcumulation factors derived from laboratory experiments have 1 imited predictive value, Moreover, in Table 3.1.1, it is seen that the biological half-thxi of I3’Cs in some biota range between approximately 40 and 100 days. Laboratory expeyirnents have riot been run for a sufficient time period in many cases for these longer-teni components to attain steady state.

3.1.3.1 Cesium Eioaccuriiulation Factors for Fishes

Bioaccumulation factors have usually been reported for flesh or whole bodies of fishes. Since concentrations of cesium and potassium do not vary greatly among organs (Nelson, 19671, flesh or whole-body bioaccumulatian factors apply to other organs as well

Potassium concentration in freshwater fishes is about 3 mg/g fresh weight and varies little among species (Kolehainen, 1972; Peterson,

1970). The range of potassium concentrations in fishes from both freshwater and marine environments is from 2.1 to 4.5 mg/I, (Fleishman, 1973; 33

Kolehmainen, 1972; Kolehmainen and Nelson, 1969). The range of potassium concentrations in freshwater and marine fishes is small considering that potassium concentration in the water range between 0.2 and 3130 ppm. Clearly, the potassium concentrations are homeostatically maintained. Thus, the bioaccumulation factor for potassium is given by Equation (2.1 -7). Since the bioaccumulation factor for potassium is given by Equation (2.1 "71, the bisaccumulation factor for cesium is given by Equation (2.1.8). Owing ts the variable availabilities of cesium relative to potassium in different environments, the value qi varies from one environment to another. Bioaccumulation factors calculated from concentrations of fallout-derived 133Cs in water and fishes collected in Finnish lakes (Kolehinainen et al., 1967; Kolehmainen et al., 1968) are given in Table 3.1.4. Within lakes, the ratio of highest bioaccumulation factor to lowest bioaccumulation factor ranges between 2.4 and 7 .I. As expected, the highest bioaccumulation factors are seen for high trophic-level fishes, namely, perch and pike, which are piscivorcs. The highest bioaccumulation factors are seen for the oligotrophic lake having a potassium concentration in water of 1 ppm. Note that the bioaccumulation factors in the oligotrophic lake are very much higher than the bioaccumulation factors in the other two lakes having the same potassium concentration but higher concentrations of organic matter. The difference appears to be related to the lower concentration of organic matter in the oligotrophic lake. Most probably, however, this difference is related to the lower suspended sol ids concentration (both organic and inorganic solids) expected in the ol igotrophic lake. The bioaccumulation factors for the fishes from the lake having a water concentration of 3.5 ppm potassium are more than 25 times lower than that of the oligotrophic lake having a mu3.. e m c3 N

01I c3 I I bII -1 I m

10I I0 I IC0 I I I

00 0 Q CUNa0 m -n Ln 7

IO0 100 IC0 iu I.. CU N .-V iI a0 0 L 0 w r- 3 N a, I1 w

.n V .r i 0 Q 0 0 L u3 +-' v 0 csl *II r--" L 0 I a, V m .r 3 L R Lt- 0 0 L m +J ;f v)

aI1 0 l a W .I V c.P- 0a L c, 0 ul 0 *P- 7 7 0 I1 0 a 35 potassium concentration of 1 ppm. If the value of qi- were constant, the bioaccumulation factors in the 3.5-ppm lake would be 3.5 tinies lower than the bioaccumulation factors in the l-ppm lake. Thus, the potassium concentration in the water alone is not very useful for predicting the bioaccumulation factors for cesiua in fishes. The conclusion that the potassium concentration in the water alone is not very useful for predicting cesium bioaccumulation factors is corroborated by the data in Table 3.1.5. The table gives bioaccumulation factors for cesium as related to potassium and suspended solids concentrations. The btoaccwqqlatfon factors for cesium in fishes were taken from Table 3.1.6. The highest bioaccumulation factors are seen for bodies of water of low turbidity. ?he bioaccumulation factors of bluegills and catfish found in the clear water body with a potassium concentration of 1.4 ppm are more than six and seven times the respective bioaccurnulation factors af bluegills and catfish from the very turbid water body with a potassium concentration of 1.3 ppm. Also, the highest bioaccumulation factor for largemouth bass is in a lake of high potassium concentration and low turbidity. Mast investigators have reported bioaccumulation factors that are cqTculated .fp~mcancentration of cesium isotopes in unfiltered water samples; however, some have reported bioaccumulation factors based on Concentration in the water after particulate material has been filtered off‘. We shall denote a “filteredtdbioaccumulation factor by BF*. Assuming a1 1 particulate material is Pi1tered out, a BF* represents an upper-1 imit Giaaccumulation factor for that and other environments of similar potassiurn concentration and sediment type since cesium on suspended sediments is to 36

a Na

0 cuQ I---

0I= *I--

h Table 3.1.6 Bioaccunulation factors for ceslum in fishes

BF t Suspended K concen- BF, BF*,~ BF*, BF, BF*, Water 501 ids Species Tissue tratipn ffJ!gt ~~~o~f~chronic stable stable treatment Location fatlout kiitl:g concentration Reference (PPrr.1 h37Cs of e37cs cesiun; cesium method ~~’~$~‘,, (PW)

____Sa 1mc t r u t ta Fb 0.4 3,900 Lake Trawsfynyd, England LowC 1.42 Percs fluviatills F 0.4 5,830 Lake Trawsfynyd, England LOW 1.42 Salino tru--tta F 0.3 3,000 English river LOW 1.42 S ;I 1ao rut ta F 0.4 1,400 English river Low 1.42 t I __Saiao trutta-- F .9 920 Engl fsh river COW 1.42 __-Sa 1no t r u tt a F 3.8 940 English river LOW 1.42 Sa 1rr.o t rutta F 4.1 260 Engl fsh river Low 1.42 Salnrcm F English river 4.1 380 Low 1.42

.5 .__--?oim,i$; aiifiu:iri; F i suo 6,400 0.7 v‘ Clinch River, Tennessee 25-185’ w - 0.7 1.36 Aplodinotus 1.3 350 4,4m u Clinch River, Tennessee 25-185 1.36 U grunniens F Roccus Chrysops 540 Clinch River, Tennessee F 1.3 8,000 0.7 25-1 85 .36 F 1.3 2,030 u Clinch River, Tennessee 25-185 rctalclruspunctatiis 150 0.7 u .35 Lepomis macroch: 140 1,700 0.7 Clinch River, Tennessee ?US F 1.3 25-185 .36 <. .. Wg i .8 310 0.7 ;I White Oak Lake, Tennessee Highh .27 i4 1.8 420 0.7 ,~ White Lake, Tennessee High 1.27 ._” Oak E 7i. 2c0 JJV g White Oak Lake, Tennessee w 1 .8 630 6.7 High 1.27 I .8 1 ao 450 320 0.7 i” White Oak Lake, Tennessee High 1.27 w il0 W i .a 2721 310 0.7 i. White Oak Lake, Tennessee High 1.27 u 1.8 240 . 530 430 0.7 p White Oak Lake, Tennessee High 1.27 Table 3.1 .6 (cont:nued)

Sljspendes SOll.?S Spec’:es Tissue tration concentratim Reference (PPl) - Mi cropterus k’ 1.8 3,59 . 9iO 620 0.7 P White Oak Lake, Tennessee High 9.27 sa 1rnoi des

Salmo qairdnerr’ F 1.7 12,030 East Twin Lake, Colorado Low 1.12

Mi cro-rus 9.6 i ,625 570 P Uintergreen Lake, Low 1.7 Xirnoider w Michfgan Perca fldvescens W 9.8 1,200 1,600 P Wintergreen Lake, Michigan Low 1.7 Hybrid sunfisa W 9.8 500 250 Uintergreen Lake, Mlchigan Low 1.7 Erirnyzon sucetta 9.8 790 300 Uintergreen Lake, Michigan Low I .l kenner; i i W

Perca W 2 . .i 2,900 Lake Maggiore, Itaiy 1.2.1.3 fluviacilis 1,100 Perca K 2.6 1,400 140 Lake Varese, Italy 1.2,1.3 w fluviati iis 03 Perce ~ 2.1 2,400 660 Lake Comabbio, Italy 1.2,1.3 fluviatilis W k: : .3 5,100 930 Lake Monate, Italj ‘I .2,1.3 Z.! Lake Maggiwe, Italy 1,330 560 l.2,’l .3 W 2.6 890 140 Lake Yarese, Italy 1.2,1.3

k 2.1 ‘1 ,000 370 Lake Comabbio, Italy 1.2,1.3 \I 1.3 1,40G 350 Lake Monate, Itaiy 1.2.1.3

I. Table 3.1.6 (continued)

Ei, BF, BF*, Suspended EF*, Water k;;J::!3 trdtioriconcen- BF;ut chronic chronic stabie treatment sol ids Species iissue (PPd yilt, faF*& $l5$qEs $'?!?& niethod Location concentration Reference T~~cs cesiua cesius (Ppm) W 2.1 790 440 Lake Maggiore, Italy 1.2,1.3 W 2.6 420 110 Lake Varese, Italy 1.2,1.3 2.1 690 420 Lake Comabbio, Italy 1.2,1.3 2,800 790 Lake Monate, Italy 1.2,1.3

1.4j South .3 F 1,230900 Par Pond, Carolina 1.3i 1.19 F 1.4 Par Pond, South Carolina 1.19 South F 1.4 1,200 Par Pond, Carolina 1.3 .I9

F 1.6 2,700 0.45 fi Lake Michigan LOW .5 F 3,300 0.45 fi Lake Michigan Low 1.6 2,400 0.45 u .5 F 1.6 Lake Michigan Low .5

F 1.6 3,400 0.45 Lake Michigan Low 1.5 F 1.500 0 , Lake Michigan 1.6 45 iO'% ? .s CI, u3 W ? .6 1,400 0.45 c Lake Michigan LO% 1.5 W 1.6 1,909 0.45 Lake Michigan Low 1.5 ii 1.6 2,100 0.45 i; Lake Michigan Low 1.5 Table 3.1.6 (conrinued)

S u s pencad K concen- BF, BF*, chronicBF, cnronicBF*, 57,^- BF*, CIater soi ids Species Tissue tration f- -out stable stable treatment ~~~~~~g Location concentration Reference (ppm) fjjcs '~~~,","L O",'ffg& ,'",'e$;:, iesjum cesium method (ppm!

Cyprinus cdrpio w 2.6 500 Carp pond 1.53 aA MF* is a bioaccunulation factor calculated from isotope concentration in water after filtra7ion.

"F = flesh.

'LOW 5 30 ppm suspended solids. d~ = piscivorous. epore s!ze of filter. fStruxiiess et a;. (1967) gN = whole fish. hHigh > 30 ppm suspended solids iA-ER anion-exchange resin. -P = 0 JPersonal comnunication.

I. some degree available to the food web. Whenever possible, we calculated both unfiltered and filtered bioacxumulation factors.

The unfi 1tered and fi1 tered bioaccurnul ation factors in Tab1 e 3.1 .S are categorized according to source of isotope: fa1 lout-derived chronic release of 137Cs, and stable cesium. In fishes frawn Italian lakes (Table 3.1.61, Bortoli et al. (1967) reported that ~i~a~~~~~~ati~~ factors calculated from distributions of fallout-derive? were on the average four times higher than bioa~curnu~a~~~nfactors calculated from stable cesium distributions. Clearly, ~~l~~~~-~er~~~d’p37Cs is somehow more available to the food web than stable cesiuna. A greater proportion of .~allout 137~s, which enters watershe s from the ~~~~~~~~r~ as a soluble radionuclide, may be in the soluble phase or on exc sites (adsorbed) of clay particles. The recently introduced lJJCs may not be in isotopic equilibrium with the stable cesium on the clay particles, especially stable cesium at non-exchangeable sites. ~~~~~~~r~~ to Eyman’s study of Wintergreen Lake, Michigan (1972) bioacc uaatinl-8 factors for

chronic release were about the sanie as ~~~~~~~~~~1~~~~~factors derived from stable ceslurn data. This may result from the tjne history of ’137Cs

as higher than during the study, We also note that the stable cesium Concentration in water was measured on water that had receive fi7 tration to remove 1srger parta”c1ss a A1 1 these patterns would suggest 42 higher--roughly by a factor of two--than the bioaccumulation factor for stable cesium. The magnitude of the difference may depend on suspended

sol ids concentrations I We have shown that the bioaccumulation factor for cesium is highly variable froni one environment to another and that much of this variation derives froni differing proportions of radiocesium relative to potassium in the soluble phase and possibly sediment type.. Unfortunately, the data necessary for accurate quantification of bioaccumulation factor relations are not available. Approximate relations, however, can be stated.

To obtain these relations we plotted the bioaccumulation factor for cesiutn in fishes against the potassium concentration in water on logarithmic graph paper, as shown in Figure 3.1.1 Unfiltered bioaccurnulation factors were used so as to better predict radionuclide concentration in the organism. For turbid environments a filtered bioaccumulation factor may be more than 10 times greater than an unfiltered bioaccumulation factor since more than

90 percent of the cesium in water nay be in the particulate phase. Fishes were grouped into two categories: piscivorous fishes and non-piscivorous fishes. By inspection we chose the relation

4 BF(CS)~= 1.5 x 10 /rKIw (3.1 .l) as an upper-bound relation for piscivoraus fishes and

BF(CS)~= 5 x 103/[K]w (3.1.2) as an upper-bound relation for non-piscivorous fishes, where [K], has units of gpm. Equations (3.1.1) and (3.1.2) are applicable to environments of low turbidity since they will greatly overpredict bioaccumulation 4.3

105

1O4

103

102

5 0.1 0.2 0.5 1 2 5 10 20 50 100 POTASSIUM CONCENTRATION IN WATER (ppm)

Figure 3.1.1 Bioaccumulation factors for 137Cs and stable cesium in fi eshwat r fishes as a function of potassium concentration in water. (All data in Tables 3.1.4 and 3.1.6 were plotted except (1) stable cesium bioaccumulation factors from studies in which 137Cs bioaccumulation factors were reported and (2) BF*s.) 44 factors for fishes from turbid environments. Dividing Equations (3-1-1) and (3.1.2) by five ga’ves rough approximations for the respective relations for fishes in turbid waters; that is,

far piscivorous fishes from turbid waters and

BF[CS)~2- 1 x 10 3/[K], (3.1 -4) fsr non-piscivorous f-ishes From turbid waters. We may consider waters having greater than 50 ppm suspended sollids as heiy turbid based on the data in Table 3.1.3. Alternatively, hioaccumulation factors for particular species may be estimated from the raw data in Tables 3.1.4 and 3.1.6. The bioaccumul ation factor selected should preferably coiw from an environment having the same suspended sol ids and potaassiim concentrations as the environment to which it is applied. Even if this is done, a degree of uncertdinty remains because of the unspecified effect 0.f sediment type ~n bioacciirnula- tion factors as well as error associated with measurement and experimental design .

3.1.3.2 Cesium Bioaccurrieslatisn Factors for Algae

As can be seen from the bioaccumulation factors for Cladophora

$omera&a and Pithsphora oedegonia (Table 3,1.7), the potassium concrn-- tration in water has little effect on the bioaccumulation factor for cesium in multicellular algae. Recently, Gertz (1973) has shown that the potassium concentration has no effect on the bioaccumulation factor for cesium in

Chlamydomonas-__1_ I-cinhardii,~ ~ a unicelluJar green alga, He found that the Table 3.1.7 Bioaccumulation factors for cesiuin in freshwater algaea

Bi oaccumul ation factors Taxon / f un cti on a 1 K concen- category t ra t ion 37cs Stable Cs Location Source of 137Cs Reference (PPI BF BF*~ BF BF* El ue-green algae: Plecotonema boryanum 3.6 1ooc Lab study 1.22 Unidentified 1.4 1,200 Par Pond, Chroni c-re ease 1.19 South Carol na Unidentified - 3 ,400 Connecticut R ve r Chronic-re ease 1.1 Diatoms: Navi cu7 a semi nul urn 14 130' Lab study 1.22 d Mixed species 1 .S 2900e Lake Michigan 1.4 P Euglenophyta: vl Euglena intermedi a 8 700 Lab study 1.51 Green algae, mu1 ti cell ul ar: Chara sp. - 20 Lab study 1.45 Chara aspera - 18 Lab study 1.45 Chara brauni i 1 60 Expe r imen ta 1 ch ann e 1 Chroni c-release 1.16 recei vi ng river water Chara fragilis - i ,000 Lake Bo1 shoe Missavo, 1.33 Russia Chara fragi is I 1 36 Lab study 1.45 Chara tornentosa - 80 Lake Bo1 shoe Missavo, 1.33 Russ i a - 120 Lab study 1.45 0.1 170 Lab study 1.48 30 59 Lab study 1.48 - 160 Lab study 1.45 8 150 Lab study 7.51 10 140 Lab study 1.57 Yougeotia sp. 9.8 3 30 500 Winterclreen Lake , Fa1 1 out 1.7 Mi chfgan Table 3.1.7 (continued)

Ei oaccumul ati on factors

Taxon/ functi onal K concen- 37cs Stable Cs Location Source of '37Cs Reference category tration (PPnl) BF 6FQb BF BF*

Green algae, mu! ti cell ul ar: Nitel la hia1 ina 170 Lake Eo1 shoe Missavo, 1.33 Russia Oedogoni ;im s p. 1,200 1.1 hectare pond Chroni c-release 1.39 Oedogoni urn sp. 3,000 Val kea1 ampi Lake, Spike input i .32 Finland Oedogoni urn vu1 gare 790 Lab study i .51 Pitho hora oedogoni a 170 Lab study 1.48 '* oedegonia 56 Lab study 1.48 Rhi zocl oni uin l,500 Lab study 1.51 P hieroglyphi cum cn Scenedesmus acumi natus 39 Lab study 1.45 Scenedesmus quadri cauda 28 Lab study 1.45 Spi rogyra communis 220 Lab study 1.51 Spirogyra crassa 28 Lab study 1.45 Spiroqyra ellipsospora 340 Lab study 1.51 Spirogyra sp. 190 Lab study 1.45 Spirogyra sp. 380 Lake Bo1 shoe Missavos 'i .33 Russia Spirogyra sp 7 50 Experi mertal channel Chroni c-re: ease 1.15 recei vinc~ river water Stigeoclonium iubricum 89 Lab study 1.22 Toli pel lopsis stel li gera 220 Lake Bo1 shoe M.issavo, 1.33 Russia Ulothrix sp. 1,400 Lab study 1.52 Ulothrix sp. 460 isere River Fa1 lout I .52 Vaucheria walzi-i 500 Expe ti menta 1 c h anne 1 Chroni c-release 1.15 receiving ri ver water Table 3.1.7 (continued)

Bioaccurnul ation factors Taxon fun c t i onal K concen- / l37CS Stable Cs Location Source of 13’Cs Reference category tra..tion (Ppm 1 BF BF*~ BF 0 F* Green algae, mu1 ticellular: Mixed species - 1,500- 4,000 Concrete-lined pond Spike input 1.39

Mi xed speci es 1.8 2 30 600f ’ White Oak Lake, Chroni c-release 1.27 ’ ,/’ .-- Tennessee Green algae, uni cell ul ar: 8 52 tab sttidy Chl amydomo~as sp. 7.5i P Chlamydomonas moewussi - 130- Lab study 1.26 Oocystis elliptica 10 670370 Lab study 1.51

Functional categories: P hy t op 1 an k ton 1.6 1900e Lake Michigan 1.4

aDry-weight bioaccumu~ationfactors were divided by 10 to convert them to wet-weight bioaccumulation factors. bA 6F* is a bioaccumulation factor calculated from isotope concentration in water after filtration. ‘Average over a1 1 temperatures. dlncludes samples where more than 80% of the cells were diatoms. eGeometric mean of bioaccumulation factors calculated from each pair of filtered (0.45 pore size) water concentrations and non-zero ”corrected” concentrations in phytoplankton given in Appendix I1 of Copelandp and Ayers (1970).

fKater ul tracentrifuged - equivalent to 0.7 p pore-size filter. 48 cesium bioaccuniiilation factor was decreased by increasing sodium concentration in water. Both Harvey (1969, 1970) and Gertz (1973) reported that the cesium bioaccumulation factor is unaffected by non-1 ethal temperatures

Bath the nutrient concentration and general health of the algae influence the cesium bioaccumulation factor far algae, since the bioaccumulation factor is higher at higher phosphate concentrations (Gertz,

1973) and the bioaccumulation factor of dead cells is lower than that of live cells (Williams, 1970; Williams, 1960), Moreover, the cesium bioaccumulation factor for mu1 t.icellular algae is higher in flowing water than in still water (Watts and Harvey, 1963). Most bioaccumulation factors for algae in Table 3.1.7 are near or 3 less than 10 I We therefore recommend that for alwe in general a bio- 3 accunrulatiin factor of 10 be used.

3.1.3.3 Cesium Bioaccumulation Factors far Vascular Aquatic Plants

Same vascular aquatic plants my obtain a significant fraction of their body content of cesium and other elements from root uptake of elements from interstial water of bottom sediments. Diffusion of elements into interstitial water from water overlying sediments is slow. This may explain why the bioaccumulation factor of fa1 lout-derived 137Cs in

Nuphar sp. (Table 3.1.8) is much lower than the corresponding bioaccumulation factor of stable cesium. Since most bioaccumulation factors in

Table 3.1.8 are less than or about lo3, we recommend that a bioaccumulation factor of lo3 be used for aquatic vascular plants. Table 3.1.8 Bioaccumulation factors for cesium in aquatic vascular plantsa

K concentration BF, llout BF, spike input BF, chronic b Reference Species in water (ppm) 15ycs of 137cs release of 137~~EF, stable Cs BF*, stable Cs Location fuli cul Azo1 la oides - Concrete-lined pond 1.22 Ceratoph.vl lum demersurn - Concrete-lined pond 1.39 Ceratophytlum sp. 9.8 370 490 Wintergreen take, 1.7 MIchigan

Elodea canadens is - ~ 1,000 Concrete-lined pond 1.39 Elodea canadensis 390 Isere River 1.52 2 Elodea canadensis 2.6 1,400 200-m carp pond 1.53 ~-Lemna minor - 500 Concrete-lined pond 1.39 Lema minor - -- _I 500 Isere Kver 1.52 Nuphar 1 uteum 3.5 130 Finnish lakes 1.30

~ Muphar luteum 1.8 41 0 Finnish lakes 1.30 Nuphar luteum 1 .o 1,000 Finnish lakes 1.30 Nuphar luteum 0.8 1 ,500 Finnish lakes 1.30 Nuphar luteum 0.6 1,100 Finnish lakes 1.30 Nuphar sp. 9.8 770 3,800 Wintergreen Lake, 1.7 Mi chi gan Nyirtphea 1utea 2.1 280 390 Lake Maggiore, Italy 1.3,1,35 - Potornogeton pectfoatus 700 Concrete-lined pond 1.39 a Dry-weight bioaccumulation factors were mu1 tiplied by 0,15 to convert them to wet-weight bioaccumulation factors. bA BF* is a bioaccumulation factor calcuiated from isotope concentration in water after filtration. 50

3.1.3.4 Cesium Bioaccumulation Factors for Eiinergent Vascular Plants

Most bioaccurnulation factors for emergent vascular plants (Table 3.1.9) are less than or near lo3; we therefore recommend that for this group in general a bioaccumulation factor of 703 be applied.

3.1.3.5 Cesium Bioaccumulation Factors for Invertebrates

There appears to be much interspecific variation in bioaccurnulation factors for crustaceans (Table 3.1 .lo). Nevertheless, most bioaccumulatian factors in Table: 3.1.10 are near lo3, therefore, a bioaccumulation 3 factor of 10 is recommended for invertebrates in general. However, bioaccumulation factor for shells of invertebrates may be much lorrrer-.

3.1.3.6 Cesium Bioaccumulation Factors for Amphibians

The few bioaccumulation factors collected for amphibians (Table 3.1.11) 4 suggest that a hioaccumulatian factor of 10 be used.

3.1.3.7 Cesium Bioaccumulation Factors for Waterfowl The few bioaccumulatian factors collected for waterfowl (Table 3,1.12) 3 suggest that a bioaccumulation factor of 3x10 be used. , .. 11 .e

Table 3.1.9 Bioaccumulation factors for cesium in emergent vascular plants

K concentration BF, llout BF, spike input BF, chroni Speci es/part f Location Reference in water (ppm) 137cs of 137cs release of 157~s ‘‘9 stable CS Equisetum fluvi ati lea 3.5 480 Finnish lakes 1.30 Equisetum fluvi atilea 1.8 7,100 , L.J Finnish lakes 1.30 Equisetum fluviatiled 1 .o 1,600 CL Finnish lakes 1 .30 I .’ Equisetum fluviatiled 0.9 1,200 ., 0% Finnish lakes 1.30 Equisetum fl uviatiled 0.8 7,400 3’‘ Fi mish 1 akes 1 .30 Phragmi tes comnunis 2.6 1,200 200-m2 carp pond 1.53 Polygoni um 1 apathi fol i urn: - 1.1 hectare pond 1.39 seeds 240 Polygonium persicaria: - Concrete-1 i ned pond 1.39 lezves 600 seeds 400 m d Sci rpus acutus : - 1.38,1.39 culms 90 Concrete-1 i ned pond seeds 400 Concrete-1 i ned pond leaves 100 Concrete-lined pond roots 400 . ,?J Concrete-1 ined pond seeds 1.1 hectare pond Scirpus americanus: Concrete-lined pond 1.39 culms seeds Typha latifoiia: Concrete-] i ned pond 1.38.1.39 leaves seeds roots

~ ~~~ ~ ~~ ~~~~- aDry-weight bioaccurrulation factors were mu1 tip1 ied by 0.20 to convert them to wet-weight bioaccumulation factors. Table 3.1 .IO Bioaccumulatjon factors for cesim in aquatic invertebrates

~~~ ~ K concen- Taxori tration BF, fallout BF* fallout BF, s ike BF, chronic chronic BF, Reference in water 137~s f37Csa Snput 737Cs release of EF*,re{y;z of stable Cs stableEF*, Cs Location (wm) 137cs Crustaceans : Garmarus iaclistris 1.7 1,000 343b East Twin Lake, 1.12,i .18 Colorado Hyallela azteca - 11,000 Concrete-lined pond 1.39 Fljlsis relicta 1.6 60' Lake Michigan 1.6 Zoop;ankton 1.6 600d LaKe M:chigan 1.4 (primarily copepods) Zoop;ankton 1.7 34ab East Twin Lake, 1.18 (Daphnl'a and Cyc;ops) Colorado Insect larvae: Chaoborus sp. 9.8 830 1,600 Wintergreen Lake, Mi chi gan Chironomid larvae 1.8 640 1,60G White Oak Lake, 1.27 Tennessee Erytharnis callocata 800 Zoncrete-; ined pocd 1.39 Ischnura sp. 800 Concrete-.i inecj pond 1.39 Snails: 2 Llninea staqnalis 2.6 1 ,3130 200-m carp pond 1.53 Rac;x japonica - 6013 Concrete4 i ned pond 1.39 53

- r. 0r. m vc

VI WL

dh Wc m

Wc -m aW W0 r 0 .-.Y x .r

WS aaJ 21:a E O rL

2 U0 Yr:

7SU 0. 0 0N .,-S s .r v) uW r 0 In 0S .r +M +J aJC 0C U0

x3 c,W -w.u OL r.L 00 7 0 VC 3 VI=L 91% 2 cwO; u mI-'wm mr: E3 54

Table 3.1.11 Biaaccumulation factors for -in amphibians (Pendl137Cs eton and Hanson , 1958)

Species/Tissue BF Bull frog tadpol e (Rana caterbei ana) : enti re 2,600 y Ut 4,500 flesh 1,000 Spadefoot toad tadpole (Scaphi opus haniinondi i n termon tw- en ti re 6,000 Bull frog adult: mus cl e 8,000 55

Table &+-.72 accurnul ati on factors for BisCs iq watepfowl '(P&s1ebn and Ffmson, 1958)

SpeciesrTissue BF American coot

(Ful ica Ia. ameri cana) : mus cl e I ,800 1 iver 2,200 bone 80 0 Common mal 1 ard (Anas pl atyrhynchos ) : muscle 2,000 liver 2,500 bone 700

Ruddy duck (Oxyura jamai censi s rubi da) : muscle 2,200 1 i ver 2,800 bone 900 56

References for Section 3.1

The numbers in the left-hand margin correspond to the reference numbers used in the tables of section 3.1.

1.1 Blanchard, R. L., and B. Kahn, The fate of radionuclides discharged from a PWR nuclear power station into a river, Symposium on Envi- ronmental Behavior of Radionucl ides Released in the Nuclear Industry, France , May 14-1 8 , 1 973 , IAEA/SM-172/26 a 1.2 Bortoli, M. de, P. Gaglione, A. Malvicini, and C. Polvani, Concen- tration factors for strontium and caesium in fish of the lakes in the region of Varese (Northern Italy) Minerva, Fisiconucleare, 324-331 (1967).

1.3 Bortali, M. de, P. Gaglione, A. Malvicini, and E. Van Der Stricht, Environmental Radioactivity Report EUR-3554e, Ispra (1 965).

1.4 Copeland, R. A., and J. C. Ayers, Trace element distribution in water, sediment, phytoplankton, zooplankton, and benthos of Lake Michigan: A base1 ine study with calculations of concentration factors and buildup of radioisotopes in the food web, Environmental Research Group Special Report No. 1, Great Lakes Research Division, Univ. Michigan, Ann Arbor, 1970.

1.5 Copeland, R. A., R. ti. Beethe, and W. Prater, Trace element distribution in Lake Michigan fish: A W.baseline study with calculations of concentration factors and equilibrium radioisotopes distribution, Environmental Research Group Special Report No. 2, Great Lakes Research Division, Univ. Michigan, Ann Arbor, March, 1973.

1.6 Edgington, D. N., E. Harvey, and M. S. Torrey, A preliiriinary study

of the chemical composition of the opossum shrimp, ..._.I.__--Mysis~ _____I-relicia, IN Radiological and Environmental Research Division Ann. Report , R. E. Rowland, ANL 7960, Part TTI, 1972.

1.7 Eyrnan, L. D., Cesium-137 and stable cesium in a hynerelatrophic lake, Ph.D. Thesis, Michian State University, East Lansing, 1972, 98 pp.

1.8 Eyman, i.D., and 3. T. Kitchings, Availability of 137Cs to fishes from ingested clays, IN Proceedings of the XIX of the International Association of Limnoloqy (SIL) , Winnipea,Congress Canada, August 22-29, 1974 (in press).

1.9 Fleishman, D, J. , Accurnula.tion of artificial radionuclide in freshwater fish, IN Radioecology, V. M. Klechkovskii, G. 6. Polikarpov, and R. M. A.leksakhin (eds.), pp. 347-369, John Ni1e.y and Sons, New Yot-k, 1973. 57

References for Section 3.1 (continued)

1.18 Friend, A. G., The aqueous behavior of strontium-85, cesium-137, zinc-65, and cobalt-60 as determined by laboratory type studies, IN Transport of radionuclides in freshwater systems, 3. H. Kornegay, W. A. Vaughn, 0. K. Jamison, and J, M. Morgan (eds.), pp I 43-60, TID-7664 (1963) *

1.11 Fujita, M., J. Iwamoto, and Kondo, Comparative of cesium and potassium in mammals--InterspeciesM. correlationmetabolism between

body weight and equilibrium level, Health Phys. .I12, 1237-1247 (1966).

1.12 Gallegos, A. F., Radiocesium kinetics in the components of a montane lake ecosystem, Ph.D. Thesis, Golorado State IJniv., Fort Collins, December 1969.

1.13 Gallegos, A. F., and F. W. Mhizker, Radiocesium retention by rainbow trout as affected by temperature and IN Radio- nuclides in ecosystems, D. J. Nelson (ed.), pp.weight, 361-371 (19731, Proceedings of the Third Natioial Symposium on Radioecology, Oak Ridge, Tennessee, May 10-12, 1971.

1.14 Gallegos, A. F., F. W. Whicker, and T. W. Hakonson, Accumulation of radiocesium in rainbow trout via a non-foodchain pathway, Fifth Annual Health Physics Society Mid-Year Topical Symposium, Idaho Falls, Idaho, 1970.

1.15 Garder, K., and 0. Skulberg, Ai? experimeqtal investi ation on the accumulation of radioisotopes by freshwater biota, Arch, Hydro- biol . -62, 50-69 (1966). 1.16 Garder, K. and 0. Skillberg, Sorption phenomena of radionuclide

to clay particles in river water, Int. J. Air Mat. Poll, I_8, 229-241 (1964).

1.17 Gertz, S. M., The fate of a radionuclide in the aquatic environment--A mathematical eqlaati nn representing ra.di oniicl ide uptake by algae, Ph.D, Thesis, Di-exel University, Philadelphia, 1192 pp,, 7973.

1 .18 Hakonson, E., A. F. Gal legor;, and F. W. Whicker, Use of cesium-133 P.and activation ana1;rsis for measurement of cesium kinetics in a montane lake, IN Radionuclides in Ecosystems, D. 3. Nelson (ed.), pp. 344-348 (1973), Proceedings of the ati onal Symposi urn on Kadioecol ogy, Oak Ridge Tennessee, May 10-12, 1971.

,/1*19 Harvey, R. S., Uptake of radionuclides by freshwater algae and fish, Health Phys, -10, 243-247 (1964). 58

References for Section 3.1 (continued)

A.20 Harvey, R. S., Uptake and loss of radionuclides by the freshwater clam Lampsilis -______radiata, Health Phys, 17, 149-154 (1969).

1.21 ~arvey,R. S., EifEcts of temperature on the sorption of radionuclides by a blue-green algae, IN Syriposiurii on Radioecology, 0. J. Nelson and F. C. Evans (eds.), pp. 256-269 (19693, Proceedings of the Second National Symposi iini on Radioecology, Ann Arbor, Michigan, May 15-17, 1967.

.22 Harvey, R. S., Temperature effects on the sorption of radionuclides by freshwater algae, Health Phys.. 19, 293-297 (1971)).

1.23 Hasanen, E., S. Kolehrtiainern, and J. K. Miettinen, Biolagical half- time of cesium-137 in three species of freshwater fish: Perch, roach and rainbow trout, IN Radioecological concentration process 8. Aberg and F. P. Hungate (eds.), pp. 921-924 (19671, Proceedings of the Iaternatianal Sympssiuni, Stockholm, April 25-29, 1966.

1.24 Kevern, N. L, Feeding rate of carp estimated by a radioisotopic method, Trans. Amer. Fisheries SOC, %(4), 363-371 (1966). 1.25 Kevern, N. R., N. A. Griffith, and T. Grimmard, Biological half- life of cesium-134 in carp ancl two aquatic insects, Health Physics Division Annual Progress Report for the Period Ending July 31, 1964, Radiation Ecology Section, pp. 101-102, ORNP- 3633 (1964).

1.26 King, S. F., Uptake and transfer OF cesium-137 by Chlam domotias, __Daphnia,.. . . ._ ...... - and bluegill finqerlings, Ecology ....45, .. 852-k

1.27 Kolrhmainen, S. E., Balances of cesium-137, stable cesium and potassiutii, in bluegill, Lepomis macrschirus Raf., and other fish in White Oak Lake, Health Phys. 23, 301-m (1972).

1.28 Kolehmainen, S., E Hasanen, ancl J. K. Miettinen, Cesium-137 levels in fish of different lirrtnologica1 types of lakes in Finland during 1363, Heal th Phys. 12, 917-922 (1966).

1.29 Kolehniainen, S., E. Ilssanen, and J- K. Miettinen, Cesium-I37 in fish, plankton, and plants in Finnish lakes, IN Radioecolayical concentration process, B. Aberg and F. P. Hungate (eds. 1, pp, 913-919 (19671, Proceedings of the International Symposium, Stockholm, April 25-29, 1966.

1.30 Kolehmainen, S., t. Hasanen, and J. K. Miettinen, Cesium-737 in the plants, plankton and fish OF the Ffnnish lakes and factors affecting its accumulation, IN Proceedings of the Fjrst Inter-

national Congress of Radi ?ti on Protection, 14 I S . Snyder (ed. ), pp. 407-415 (1368), Pcmagnn Press, Oxford, 1.31 ~a~e~~ai~en~S. E., and D. 3. Nelson, Balances of cesium-137, stable cesium, and the feeding rates of bluegill (Le om-is mcrochirus.___--- Raf. 1 in White Oak Lake, ORl\riL.-444Ei ~~~z$T--

1.32 ~~~~~~ain~n~S., Komantschuk, :I* Takatalo, and J. K. Miettinen, Pol 1 uta”an experiments idi th Cs-137 in I akes of two 1a’rr!nsl ogi cal ly different types e IN CongrGs irternational sur la a radio-protection du milieu devant le developpement des utilisations s.tacifiques de l’energie nucleare, pp. 181-200, Tulouse, 14-16 Wars 1967, Societe Francaise de Radio-protection, Paris (~~~~~ 1 -33 Kul i kov N V. , Kadj oecol agy of‘ freshwater plants and animals IN ~adi~~colo~y~V, V. Klechkobskii, G, G. Polikarpov, and I?. P4. Aleksakhin (eds.), pp. 323-337 (19731, Wiley, New York, 1.34 Marshall, J. S., and J, Ma LeRcy$ Iran, manganese, cobalt, and zinc cycles in a South Carolina reservoit,, EN i?adionnc‘iidpr, in ecosystems, 0, J. Nelson led.) fi pp. 465-473 (1973) 5 Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee2 May 10-42, 1971

1.35 ivlerlini, M., F. Giradi, and I;. Pozzi, Activation analysis .in studies of an aauatic ecosystem, IN Nuclear activation techniques in the 1 ife sciences pp. 615-629 (1967) Proceedings of the Symposium held by the IAEA in Anisterdarn, May 8-12, 1967, also /I%, SM-91/7,

1.36 Nelson, D. J., Cesium, cesium-137, and potassium concentrations -in white crapPie and other Clinch River fish, IN Synrpssiium on radioecology, D. J. Nelson and F. C. Evans ( s * ), pp. 240-248 (1969) Proceedings of the Second National S posiirni ow Radio- ecology, Ann Arbor, ~~ic~i~a~,Play 15-17, 1967.

1.37 Nelson, J“, N. A. Griffith, and S. A, Kucker, Radionuclide excretionD. studies IN Wealth Physics Division Annual Progress Report for Period Ending July 31, 1969, pp. 119-?20, r)RNk-4446 (197Q) - 1.38 ton^ R. C., Effects of solne environmental factors on bi’o- aceuniul ation of cesi urn-1 37 in ai aquati c communi “cy, J Biology Research Annual Report for 1958, 3. 9. Davis pp. 42-46 (1959).

R .I’’’g;’~~.~e~~~-~’~~~,and @.Hanjon, Absorption of cesium-137 components W. \ by of an aquatic cominunity, IN Proceedings of the! Second Iwtern~~i~na~Conference on eaceful Uses of Atomic Energy, Vol, -18, 419-422 (195&), U. Geneva, September 1-3, ‘i 1958, References for Section 3.1 (continued)

1.40 Pendleton, K. C., C. W. Mays, R. D. Lloyd, and B. W. Church, A trophic level effect or1 cesium-137 concentration, Health Phys. -11, 1503-1510 (1965).

1.41 Peterson, H. I., Jr., Stable element and radionuclide recon- centration in aquatic systems. I. Stable element effects on equil ibriiirn situation , IN Proceedings of the Fifth Annual Health Physics Society Midyear Topical Symposium, Health

Physics Aspects of Nuclear Siting, Vol a 2, pp. 289-328, AEC CONF 701106 (1970).

1.42 Preston, A., 0. F. Jefferies, and J. !d. R. Dutton, The concentration of cesiurn-137 and strontium-90 in the flesh of brown trout taken from rivers and lakes in the British Isles between 1961-1955, The variables determining the concentrations and their use in radiological assessment, Water Res. 475-49s (1967). 1,

1.43 Reynolds, T. D,, and E. F. Gloyna, Uptake and release of radionuclides by stream sediments, IN Proceedings of the Second Inter- national Conference on Advances in Water Pol 1 ution Research, Vole 9 , 6. Jaag (ed, ) , pp. 151 -1 72 (1964), Pet-qarnon Press, New York.

1.44 Struxness, E. G,, P. H. Carrigan, Jr., M. A. Churchill, K. E. Cowser, R. J. Morton, D. J. Nelson, and F. L. Parker, Compre- hensive report of the Clinch River study, U.S. AEC Report BRNL- 4035, Oak Ridge National Lalsoratot-y , Oak Ridge, Tennessee, 1967.

1.45 Timofeyeva-Resavskaya, Ye. A., Distribution of radioisotopes in the main components of freshwater bodies (Monograph), Tr-. Inst. Biol. Akad. Nauk SSSR, Ural’skiy filia 30,...... 1-78 (JPRS 21.816) 1963.

1.46 Vanderploeg, ti. A., R. S. Booth, and F. H. Clark, A specific activity and concentration model applied to cesium movement in an oligotrophic lake, IN Mineral Cycling in Southeastern Ecosystems, F. G, Howell (ed.), (in press), Symposium held 1-3 Flay 1974, Augusta , Georgia. /” .jl .47 Watts, J. R., and R. S. Harvey, Uptake and retention of 137Cs by a blue-green alga in continuous flow and batch culture systems,

Lirnnsl. Oceanogr. I8, 45-49 (7963).

1.48 Williams, L. G., Concentration of strontiiim-85 arid cesium-137 from \vat& sol utions by C1 adaphclra and Pi thophora J. Phycol . -6, 314-316 (1990).

1.49 Williams, L, G., Uptake of cesiurn-137 by cells and detritus of Eugl ena and Chl ore1 1a, Lirrinol . Oceansgr . 5, 301 -31 1 (1960). References for S,ection 3,l (continued)

1.50 Williams, L. G., and Q. Pickeeing, Direct and food-chain uptake of cesiurn-137 and strontium-85 in bluegill fingerlings, Ecology -42, 205-206 (1961).

1 ,1,:51 Williams, L. G., and H. D. Swnnson, Concentration of cesium-137

\/” by algae, Science _I127, 137-1813 (1958). 1.52 Wlsdek, S., The behavior of cesium-137 in freshwater reservoirs with particular reference to the reactions of plant communities to contamination, IN Radioecological concentration process, €3. Aberg and F. P. Hungate (eds. ) , pp. 897-912 (1967), Proceedings

of the International Symposium, Stockholm, April 25-29, 1971 ~

1.53 Wlodek, S., M. Bysiek, and D, Grzyboska, Autochthonous caesium and behavior of allochthonous caesium in a fish pond, Ecologia Pslska, Seria A, Vol. -18(30), 1-11 (1970).

3.2 STRONTIUM

3.2.1 Stronti um Metabol i sm

As a result of their chemical similarities and the greater abun of calcium in nature, calcium is a non-isotopic carrier for strontium. Since calcium is homeostatical ly maintained at constant concentrations in fishes (Agnedal, 1967; Ophel and Judd, 1973; Peterson, 1970; Reed and Ne1 son, 1973; Suzuki et a9 . , 1972; Templeton and Brown, 1964) and presumably other animals the bioaccumulation factor for calcium in animals is given by Equation (2.1.7). The bioaccumvlation factor for strontium is then given by Equation (2.1.8). These relatjons do not hold for algae since the calcium concentration ii! water does riot appear t~ greatly influence the bioaccumulation factor for strontium in aigile (Kevern, 1964a; Williams, 1970). The primary uptake of calcium and strontium in fishes, and probably in many aquatic invertebrates, occurs directly from the water. The gill membranes of fishes are the primary sites of ca’lcium and strontium uptake (Nelson, 1966; Suzuki et al., 1972; Templeton and Brown, 1964). Only about one-tenth of the calcium and strontium taken up by fishes is through the food chain (Agnedal, 1966). Since uptake from water is the pra’mary pcthway, the food chain dynamics of calcium and strontium are of little importance in the determination af bioaccumulation factors for strontium in aquatic organisms. Thus, the discrimination coefficient qi in Equation (2.1.8) is not a function of trophic level in fishes because the primary source of strontium to fishes is water. The discrimination coefficient has also been shown to 64 be independent of the calcium concentration in water (Rgnedal, 1967;

Ophel and Judd, 1973; PetersonS 1970; Reed and Nelson, 1973; Suzuki et a1 . , 1932; TmipSeton and Bro~n,1964).

3.2.2 Envi ronnien $a 1 S tron t i um

Strontium is commonly found in nature with calcium, its closest chemical analogue (Nelson, 1961). Strontiutn concentrations in freshwater rivers and lakes may range between 0.004 and 0.23 ppm (Nelson, 1467; Ophel et al., 1972; Suzuki et al., 1972; Templeton and Brov~n,

1964). The more abundant element calcium is found in freshwater rivers and lakes in concentrations ranging Prom 1.9 to 114 ppm (Austin, 1963;

Ben-a’nson et al., 1966; Nelson, 1967; Ophel et al., 1972; Suzuki et al., 7972; Templetan and Brown , 1964). Strontium has physicochemical properties similar to calcr’um and, like calcium, appears mainly in ionic form in water (Templeton and Brown, 1964). Strontium and calcium are not strongly sorbed by suspended particulate materials in the water.

The fraction of strontium present in the suspended phase is given far a number of freshwater ecosystems in Table 3.2.1. The fraction of strontium in the suspended phase ranges between 0.9 and 10%. Since qi is independent of [CaIw and because both calcium and strontium have nearly the same physicochemical forms in water, the discrimination coefficient is not expected to vary much among sites,

3.2.3 Review of Strontium Bioaccumulation Factors

Bioaccumul ation factors are di scwssed for the fol 1owing categories of aquatic biota: 65

Table 3.2.1 Percentage of radiostrontium in water in the particulate phase in some freshwater ecosystems

Mean % in Suspended Water part! cul ate solids Separation Reference body method phase ( ppm) White Oak Lake, 2 31 (mean) U1 tracentri fu- 2.24 Tennessee& gation: Par- ticles >0.7 1-1 i nch Ri ver, 2-9 25-185 U1 tracentri fu- 2.24 C1 Tennesseea gation: Par- ticles 4.7 1-1 Tennessee River, 9-10 9-22 U1 tracentrifu- 2.24 Tennesseea gation: Par- ticles >0.7 p Experimental Pond b 0.9 21 0.45-p fi1 ter 2.7 aRange of mean of three river statims. bBased on five water values sampled between 24 and 80 days after introduction of radi onucl ides. 66

1. Fishes 2, Algae

3" Aquati c and Emergent Macrophytes 4. Molluscs.

I-/Re biaaccurnulation factors were examined for taxa in each of the above categories. For a1 1 taxa except algae, the strontium bioaccumulation factors were derived only from concentrations in natural or semi -natural bodies of water.

A regression analysis of In BF(Ca)i versus In [Ca], for brown trout (Templeton and Bt-own, 1964) and for three species of fish from Swedish lakes (Agnebal, 1955) yielded a linear plot with a slope of appvoximately -1 and a high degree of negative correlation (Petersoil,

1970). Tile bioaccumulation Factor for calcium can thus be given by tquakion (2-1.7). Since calcium is a nnn-isotopic carrier for stronti uiii, the bioaccumul at3m factor for stronti um is gi ven by tquation (2,l.B). The analyses by Peters~n(1970) of data from

Tetripletan and Brown (1964) and Agnedal (1951) clearly showed that the qi values for freshwater fishes arc independent OF the concentrations of calciuin in the water. This imp1 ies that the regression of ln BF(Sr)i on lti [Ca]bv has a slope of about -1.

The stw-onti uni bioaccumul ation factors, cal ci um concentrations in fishes and water, and qi values for strontium in fish bone and flesh are given in Table 3.2.2 for studies that reported calciuni concentration in water. Only data based on several measurements of strontium in fishes and water are listed. Some stiadit"~have 67

1 isted values for the strontium d-iscrimination coefficients in bone and no ~ioacc~mulationfactors. En these studies, the discrimination coefficients were calculated from strontium and calcium concentrations

in ashed bone. The ratio between ash weight and fresh weig is not constant for all species or for all ashing procedures. Since a fresh weight cannot be assumed for a given ashed weight, bioaccumulation factors for strontium could not be calculated from these studies. Because qi is not a function of feeding habits, we dec-lded to derive a relation between the struntium bioaccumulation factor and

[Ca], for all fish data combined. To do this, regressions of In

BF(Srli versus In [Ca], for all fishes combined were made. Results of these regressions for fish flesh and bone are given in Table 3.2.3 and in Figure 3.2.1. The slope of the regression for fish fles not significantly different (P C.05) from the expected slope of -1.

The slope of the regression for bcme, a1 tho~ghsignificantly different from -1 (P c O.OS>, a’s not apprreciab y different, To predict the bioaccumulation factor of strontium in fishes we recommend use of Figure 3.2.1 or the relation given in Table 3.2.3.

3.2.3.2 Strontium ~i~~cc~m~la~~~~Factors for Algae

The ~ioa~~u~u~at~~nfactors for strontium in algae are given in

Table 3.2.4 for both field and laboratory studies. In a labarator-y study, Kevern (1964) showed that the bi~~ccu~ul~tio~factor for s~r~nt~~~~in the unicellular green alga Oacytis eremosphaeria was independent of the calcium concentration in the water. In Table

3.2,4, the multicellular algae -Pithophora and ---Cla hora CJ ~-.--omwata

(Williams, 19709 show only a small decrease in the ~~~~~~~~~~ati~~ factor for strontium accompanying a very large increase in calciuin Table 3.2.2 8ioaccumu~ationfactors and discrimination coefficients for s-tront-ium in fishes

Ca Concentration Species Location Reference Tissue Water BF(SrIi qi

Perch Bone 231 03 2.0 4 $30 0.35 Lake Langsjon, 2.7 Musc’l e 480 92 0.38 Sweden Perch Bone 31000 40.0 170 0.22 Lake Storacksen, 2.1 Sweden Perch Bone 31 600 46.0 160 0.23 Lake Erken, Sweden 2.1 Muscle 390 3 0.35 Perch Bone 43600 54.0 130 0.16 Lake G1 isstjarn 2.1 Muscle 430 1.3 0.16 Sweden Perch Bone 39400 4.4 1,840 0.21 Cake 111 kesjon, 2.1 Sweden Perch Bone 40000 12 41 0 0.12 take Viggen, Sweden 2.1 Perch Bone 34200 18 330 0.17 Lake Or1 angen , Sweden 2.1 Perch BOW 39800 26 380 0.25 Lake Magel , 2.3 Sweden ungen Perch Bone 31 700 63 100 0.20 Lake Sovdesjon , 2.1 Sweden Perch Bone 302OcI 03 110 0.23 Lake Nittsjon, Sweden 2.1 Pike Bone 42500 2 .0 4,930 0.23 Lake Langsjon 2.1 Muscle 91 0 125 0.27 Sweden 69

I-- - N N N

t a, c a

'zs,a, v)

n % a S c Ln 0 n '3 aJ0 c m r rg b)aJ YKal 7 0 I-w 3-3 0L CL (u a3 Ya, 2 U) Ya! rg rcl -1a -1 -J

0 N W Ln N F W a 0

0 0 Oh om l-- Cna Lo I--- mN 00, .)I u- co N

e2 d Ev 6 -co

0 C 0 0 0 0 II Lo N Ln b- 0 b- -LD

Jza, a, c ta 0 0 a m M a

Q) Y Ya YaJ *l- .F *I- a Q n 70

r-h

h 0 cr)i

El 0

1 0 cu N

0 *F mt c, c 0

Ya, M --1 -!= V L a, a

s 0 0 m r- N

h Ln m cn

IIaJ 0 f

F v) -.I-- .A 51 c,m m 'I-- >

n 72

a, Su 7 c LW N N aJ Lt- N (u CY

0 -I- L fa t, 611:

n Ya a J -c V L W a

.-N 0

m m Lo

cW 0 1 M

v) "T- E 0 aW J 73

r"- N cu

0 'r- SI m c, 0r; 0)a n n 0 a t cu 0 ia m L3 sc 0 t: '-I-

co h 0

cu Q F

N ca, V 0 M 3aJ c or- c, c 0 V U cu N m

P-aJ ala b- 74

2 42

0E t--20 2 0 402 !- a

5 0.4 0.2 0.5 1 2 5 40 20 50 100 CALCIUM CONCENTRATION IN WATER (pprn)

Figure 3.2.1 Bioaccurnulation factors for strontium in freshwater fishes as a function of calcium concentration in water. Table 3.2.3. Parameters from linear regressi'ons between In BF(Sr)i and In [Calw for bone and flesh of fishes. [To predict BF(Sr)i use the relation BF(Sr)i = exp(intercept + slope x 111 [Ca],] .I

Tissue Intercept ~f:SE Slope i- SE R2

Eone 9.59 t 0.18 -1.15 + 0.06 0.901

F1 esh 5.18 ?1 ?.I1 -1.2'1 5 0.37 0.736 76 Table 3.2.4 (continued)

Ca concentration Spec ies Location Reference in water (ppm) BF(Sr)i qi Sp irogyra s p . 5.7 120 Perch Lake, Ontario 2.20 Spi rogyra sp 5.4 9 00 Experimental Channel 2.8 Stigeocl oni urn 1ubricum 10.5 12OC Labo rato ry 2.12 Vaucheria wal zii 5.4 1,400 Experimental Channel 2.8 aDry-weight bioaccumulation factors were divided by 10 to convert them to wet-weight bfoaccumulation factors.

%ersonal communication.

'Average over a1 1 temperatures. concentration in the water. Thus it may be concluded that the bioaccumulation factor for strontium in algae is relatively independent of the calcium concentration in the water. Until additional data 3 are available, a strontium bioaccumulation factor af 2 x 10 is recommended far algae.

3.2.3.3 Strontium Bioaccumulation Factors for Aquatic and Emergent Macrophytes

The bioaccumulation factors for aquatic and emergent macrophytes are found in Table 3.2.5. Discrimination coefficients are given when available The data do not permit us to infer whether the strontium bioaccumu ation factor is dependent on the calcium concentration in

the water On the basis of these few data, we recommend a bioaccumu- 2 lation factor of 2 x 10 for aquatic and emergent macrophytes.

3.2.3-4 Strontium Bioaccumulation Factors for Molluscs

Table 3.2.6 gives the experimental data obtained on bioaccumu-

lation of strontium in mollusc shells. The data in Table 3.2.6 and Equation (2.1.8) were used to calculate a bioaccumulation factor

relation for strontium in mollusc shells. Mean values of qi, the

discrimination coefficient, and xi*, the concentration of calcium

in mollusc shells, are taken from Table 3,2.6 to derive the relation. For mollusc shells, the strontium bioaccumulation factor relation

derived was

6.8 x 104 RF(Sr)i = ~a-~-- where [ca], has units of gym. I' .

Table 3.2.5 Bioaccumulation factors and discrimination coefficients for strontium in aquatic and emergent macrophytes

Ca Species concentration BF(SrIi qi Loca ti on Reference in water ( ppm) ., Braseria schreberi 6.3 0.55 Perch Lake, Ontario 2.21 Ceratophyll um demersum 5.7 220 Perch Lake, Ontario 2.20 El odea canadensi s 23 1.2 Lough Neagh, U.K. 2.26 Fontinal is sp. 6.0 240 1.2 Perch Lake, Ontario 2.20, 2.21 Myriophyl lum spicatum 2.4 1.2 Devoke Water, U.K. 2.26 Myriophyl lum spicatum 23 1.5 Lough Neagh, U.K. 2.26

Nu p har va~ r i ega t-~ um 6.3 1.8 Perch Lake, Ontario 2.21 Nyniphea odorata 6.3 0.90 Perch Lake, Ontario 2.21 Pontederia -~cordata 6.3 1.4 Perch Lake, Ontario 2.21 Potamogeton natans 2.4 1.5 Devoke Water, U.K. 2.26 Potamogeton pecti na tus 23 1.7 Lough Neagh, U.K. 2.26 Potamogeton perfol iatus 23 0.84 Lough Neagh, U.K. 2.26 Potamogeton pusillus 6.0 I90 2.0 Perch Lake, Ontario 2.20, 2.21 Scirpus fl uitans 2.4 0.57 Dover Water, U.K. 2.26 Scirpus subterminal is 5.7 /.' , 301:;' Perch Lake, Ontario 2.20 LL.4 *' Sparangium fl uctuans 5.7 150 Perch Lake, Ontario 2.20 Sparangi um sp . 2.4 0.89 Devoke Itlater, U.K. 2.26 U tri cu 1 aria vu7 ga rl' s 5.7 9 60 Perch bake, Ontario 2.28 80

n C.n

NLnln 03 I----- P- omw 0

hhh NNeU

ebb-a00 000 Table 3.2.6 (continued)

~ ~~ Ca Concentration Location Reference Species Shell Water BF(SrIi qi (g/d (PPd Union inae (con t 'd)

Ambl ema costa ta 0.40 19 3,202 0 a 15 Tennessee River, 2.16 Tennessee Megalonaias gigantea 0.40 19 2,995 0.14 Tennessee River, 2.16 Tennessee Cycloraias tuberculata 0.40 19 3,235 0.15 Tennessee River, 2.16 Tennessee Lamsil inae Plagiola lineolata 0.40 19 3,170 0.15 Tennessee River, 2.16 Tennessee

Mean Ci 0.40 Mean qi 0.17 82

There is not a great deal of experimental data available on bioaccuriu- latian of ,strontium in the soft tissue of niolluscs. A recommended bioaccumulation factor for strsntium in the soft tissue OF molluscs made the of the experimental data that are was on basfs limited available @rungs, 1965; Harvey, 1969; Merlini et al., 1967; Nelson, unpublished data; Ophel , 19631. The recommended strontium bioacctiniu- lation factor for the soft tissue? af malluscr is 3 x lo2. It rnus t be woted however that this value is given for bodies ~f water ~ith calcium concentrations of approximately 20 to 30 ppm. The bioacczrmu- lation factor is of CQ~IPS~a funct-isn of the calcium and strontiuin canccntration in the particular body of water and bisaccumulatior factors as high as 720 were Pound for waters of very low calcium cantent [Ophel , 1963). 83

References for Section 3.2

The numbers in the left-hard margin correspond to the reference numbers used in the tables of section 3.2.

Agnedal, P. O., Calcium ar;d strontium in Swedish waters and fish and accumulation of strantium-90, IN Radioecological concentration process, 8. Aberg and F. P. Hungate (eds.), pp. 879- 896 (19671, Proceedings of the International Symposium, Stockholm, April 25-29, 15‘66.

Austin, J. H., Strontium uptake by algae, IN Transport of radionuclides in freshwater systems, B. H. Kornegay, W. A. Vaughan, D. K. Jamison, and J. M. Morgan (eds. ) , pp. 231-245, TID-7664 ( 1963).

Beninson, D. E., E. van der Elst, and D. Cancio, Biological aspects in the disposal of fission products into surface waters, IN Disposal of radioactive wastes into seas, oceans, and surface waters, pp. 337-354, Proceedings of International Atomic Energy Agency Symposium, Vienna, 1966, STI/PUB/126.

Bortoli, M. de, P. Gaglione, A. Malvicini, and C. Polvani, Concentration factors for strontium and caesium in fish of the lakes in the region of Varese (Northern Italy), Minerva Fi siconucl eare 324-331 ( lC167 1 - Bortoli, M, de, P. Gaglione, A, Malvicini, and E. van der Strieht, Environmental Radioactivity Report, EUR-3554e, Ispra (1966) . Brungs, A,, Experimental uptake of strontiurn-85 freshwater organisms,W. Wealth Phys. -11 , 41-46 (1965). by Friend, A. G., The aqueous behavior of strontium-85, cesium- 137, zinc-65, and cobalt40 as determined by laboratory type studies, IN Transport of radionuclides in freshwater systems, B. H. Kornegay, W. A. Vaughn, D. K. Jamison, and J. M. Morgan (eds. ) , pp. 43-60 , TID-7664 (1 963). Gardner, K., and 0. Skulberg, An experimental investigation on the accumulation of radioisotopes by freshwater biota, Arch. Hydrobiol . -62, 50-69 (1966). Harvey, R. S., Uptake of radionuclides by freshwater algae and fishh,Health Phys. -lo., 243-247 (1964). 84

3.2 ,--, References for Section (continued)

i 2-10;Harvey, R. S., Effects of temperature on the sorption of T 1 radionuclides by a blue-green algae, IN Symposium on Radioecology, D. J. Nelson and F. C. Evans (e&.>, pp. 2GG- 269 (1969), Proceedings of the Second National Symposium on Radioecology, Ann Arbor, Michigan, May 15-17, 1967. 2.11 Harvey, R. S., Uptake and loss of radionuclides by the freshwater clam Lam si1 is radiata (Gmel . ) , Health Phys. -17, 149-1 54 (1 96+ --II_ f - %’\ ’ 2.121 Harvey, R. S., Temperature effects on the sorption of radio- \ / nuclides by freshwater algae, Health Phys. 19, 293-297 (1970). (2.13 kvern, N. R. , Strontium and calcium uptake by the green \ - -. - lalgaI Oocystis eremosphaeria, Science 145, 1445-1446 (1964) 2.14 Kevern, N. R., N. A. Griffith, G. M. Rosenthal, Jr., and D. J. Nelson, Uptake and accumulation of strontium and calcium by algae and snails, Health Physics Division Annual Progress Report for the Period Ending July 31, 1964, Radiation Ecology Section, pp. 96-101, ORNL-3697 (1964a). 2.15 Merlini, M., F. Girardi, and G. Pozzi, Activation analysis in studies of an aquatic ecosystem, IN Nuclear activation techniques in the life sciences, pp. 615-629 (1967), Proceedings of the Symposium held by the IAEA in Amsterdam, May 8-12, 1967, also STI/PUB/155, SM-91/7. 2.16 Nelson, D. J., The strontium and calcium re?alionships in C1 inch and Tennessee River mol 1uscs IN Radioecology , V. Schultz and A. W. Klement, Jr. (eds.), pp. 203-211 (1963), Proceedings of the First Natiana? Symposium on Radioecology, Fort Collins, Colorado, September 10-15, 1961. 2.17 Nelson, D. J., The prediction of strontium-90 uptake in fish using data on specific activities and biological half-lives, IN Radioecological concentration processes, B. Aberg and F. P. Hungate (eds. ) , pp. 843-858 (1967), Proceedings of the International Symposium, Stockholm, April 25-29, 1966. 2,118 Nelson, D. J., Ecological behavior of radionuclides in the Clinch River and Tennessee River, IN Proceedings af Reservoir Fishery Resources Symposium, pp. 169-187, Athens, Georgia, April 5-7, 1967. 85

References for Section 3.2 (continued)

2.19 Ophel, I. L., The fate of radiostrontium in a freshwater community, IN Radioecology, V. Schultz and A. W. Klernents, Jr. (eds.), pp. 213-216, Reinhold, New York (1963).

2.20 Ophel, I. L., C. D. Fraser, and J. Judd, Strontium concentration factors in b-'ota and M.bottom sediments of a freshwater lake, EUR-4800 (May 1972).

2.21 Ophel, I. L., and J. M. Judd, Strontium-calcium relationships in aquatic food chains, IN Radionuclides in ecosystems, D. 3. Nelson (ed.), pp. '221-225 (1973) Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee, May 10-12, 1971. 2.22 Peterson, H. T., Jr., Stable element and radionuclide recon- centration in aquatic ecosystems. I. Stable element effects in the equilibrium situation, Fifth Annual Health Physics Society Midyear Topical Symposium, Idaho Falls, Idaho, November 3-6, 1970.

2.23 Reed, 3. R., and D. J. Nelson, Radiostrontium uptake in blood and flesh in bluegills (&omis macrochirus), IN Radionuclides in ecosystems, D. J. Nelson(ea.Tpp. 226-233 (19731, Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee, May 10-112, 1971.

2.24 Struxness, E. G., P. H. Carrigan, M. A. Churchjll, K. E. Cowser, R. 9. Morton, 0. Js Nelson, and F. L. Parker, Comprehensive report of the C1 inch River Study, ORNL-4035

(April 1967) e

2.25 Suzuki, Y., R. Nakamura, ard T. Ueda, Accumulation of strontium and calcium in freshwater fishes of Japan, Japan Radiat. Res. SOC. -13, 199-207 (1972).

2.26 Templeton, W. L., and M. Brown, The relationship between concentrations of calcium,V. strontium and struntium-90 in wild brown trout, Salmo trutta L., and the concentrations of the stable elemxisome waters of the United Kingdom, and the implications in radiological health studies, Int. 3.

Air Water Poll. I8, 49-75 (1964).

2.27 Timofeeva, A., and N. V. The role of freshwater plants in theN. accumulation ofKul~kov, 90Sr and its distribution over the components of a reservoir, IN Radioecological concentration processes, Aberg and F. P. Hungate (eds.), pp. 835-841 (1967) ProceedingsB. of the International Symposium, Stockholm, April 25-29, 1966. 86

References for Section 3.2 (continued)

2.28 Williams, L. G., Concentration of strontium-85 and cesium-137 from water solutions by Cladophora and Pithophora, J. Phycol. -6, 314-316 (1970). 3.3 TRITIUM

Tritium was included in this report because of a concern that the relative kinetics of tritium and protium resulting from tritium’s heavier mass (”isotope effect”) might lead to a preferential accumulation of tritium over protium. Experiments in aquatic systems indicate that this does not occur and that the bioaccumulation factor for tritium is less than or about equal to the bioaccumulation factor for protium, which has a bioaccumulation factor approximately equal to 1. This section discusses the movement and distribution of protium in aquatic organisms and reviews experiments that compare the behavior of tritium and protium in aquatic systems. In the discussion the term hydrogen is used to refer to both isotopes of the element collectively.

3.3.1 Hydrogen Bioaccumul ation

Hydrogen atoms in living organisms can be separated into two major categories or pools. The first pool, tissue-water hydrogen (TWW), is defined as all hydrogen atoms preseqt in water molecules within the organism. The second pool , ti ssue-bound hydrogen (TBH 1 is defined as all hydrogen present in organic molecules, such as proteins, fats, and carbohydrates. As an illustration, a fish consists of approximately 75% water and 25% dry tissue. The percentages of water and dry tissue, the relative sizes of the two hydro- however, do not accurately reflect gen pools because on a weight basis water contains 11% hydrogen, while dry fish muscle contains 8%. As a Pesult the tissue-water hydrogen constitutes 8Q% of the total fish h:jdrogen and the tissue-bound hydrogen constitutes 2Q%- Because of this lower hydrogen concentration in dry 88 tissue, the bioaccumulation factor for protium in fish is approximately

0.93. Subsequent discussions focus on tritium and the degree to which

its heavier mass may cause it 'to have a different bioaccumulation factor

from that 0.f protium.

3.3.2 Pathways of i-iydrogen En try into Aquatic Organi sms

Tritium enters aquatic systenis in the form of tritiated water,

that is, HTO, where T represents tritium; H, protiuni; and 0, oxygen. Tritiated water behaves like HOW and mixes rapidly with the tissue water of aquatic organisms. From the tissue water pool tritium can enter hydrogen sites in organic tissue. These hydrogens comprise the

tissue-bound hydrogen pool which may be further subdi vi ded into two components termed exchangeable (ET) and nonexchangeable (NET) , which have different rates of turnover and different modes of uptake and

lass (Figure 3.3.1). The exchangeable component consists of those

hydrogens banded to OH, COOH, NH, SH, and ortho and para positions

of phenol groups. These hydrogens undergo a relatively rapid chemical

exchange reaction with hydrogens of water molecules in the tissue

water. This exchange is not dependent un enzymatic reactions and occurs in melabolically inactive tissues, such as wood, as well as metabu1 ical ly active tissue. The nonexchangeable component consists

predominantly of hydrogens bonded to carbon atoms of organic tissue. Movement of hydrogen into and aut of this component is dependent on various enzymati :: reductions and oxi dat-ians of tissue organics . Various metabolic reactions reduce organic molecules by incorporating hydrogen from tissue water into stable carban-hydrogen hands in ORNL- DWG 74-4996R2

NON EXCHANGEABLE FOOD TISSUE .BOUND

TISSUE WATER EXCHANGEABLE TISSUE BOUND

3.3.1 Figure Pathways of hydrogen entry into the hydrogen components of an aauatic animal. 90 i nterrnedi ary products used to synthesize lipids and proteins . Hydrogen is iberated from this nonexchangeable poo by oxidation reactions, These oxidations enzymatically remove nonexchangeable hydrogen from tissue organics and ul kimately combine the hydrogens - with oxygen to form tissue water. Ihe turnover of this pool is much slower than the exchangeable pool. For a discussion of reactions that incorporate hydrogen from tissue water into the nonexchangeable pool , see Smith and Taylor (1969).

A second pathway for hydrogen entry into the nonexchangeable component, which applies to animals, consists of the ingestion and

incorporation of intact food molecules containing nonexchangeable hydrogen. The absence of this pathway in plant uptake and its importance in animals and food chain transfers is discussed in Section 3.3.4.

3.3-3 Compartment Size and Turnover Rates

Tissue-water components constitute from 80 to 93% of aquatic

plant weight and averages 75% in aquatic vertebrates. Tritium has been shown to move rapidly from ambient water into tissue water, with bioaccumulation factors approximating unity and biological half-times measured in minutes for small unicellular algae to hours for large aquatic macrophytes and aquatic vertebrates (Elwaod, 1973; Harrison and Koranda, 1973; Stewart et al., 1973). Tritium in emergent portions of rooted macrophytes, such as cattails, may fail

to reach steady-state with ambient water possibly because of a mixing of water in leaves with less contaminated moisture in the air (Raney and Vaadia, 1965). 91

Relative sizes of exchangeable and nonexchangeable components have not been determined for aquatic organisms. Data from small mammal experiments, indicating t'rlat exchangeable hydrogen constitutes 30% and nonexchangeable hydrogen 70% of the tissue bound component, were used as tissue-bound compartment sizes in constructing Table

3.3.1 and Figure 3.3.1 (Pinson and Langham, 1957; Siri and Evers, 1961).

Data on turnover rates of tit-itium in the bound components of aquatic organisms are sparse Elwood (1 973) and Patzer (1973) working on goldfish and mosquito fish, respectively, demonstrated 8-day half-times for tritium elinination from the combined tissue- bound component. For aquatic snails, a 62-hr half-time for 70% of the tissue-bound component has been demonstrated (Stewart et a1 ', 1973). A typical study on rats demonstrated half-times of 22 days for 50% of the tissue-bound tritium and 130 days far the other 502

(Thompson and Ba?lou, 1956). Half-times in humans characteristi ea1 ly show a short tissue-bound componEnt of 30 days and a longer component of 300 days (Pinson and Langham, 1957). In one case Pinson and Langham observed a 2020 day half-time in a chronically exposed human. A1 though aquatic organisms have demonstrated longer half-times for tissue-bound tritium than for tissue-water tritium, no measurements of elimination rates for exchangeable and nonexchangeable tri ti urn have been reported for aquatic organisms exposed to tritium in their food and water under control led conditions. Results of experiments cited above indicate some of the tissue-bound components may require a significant fraction of the life of the organism to equilibrate 92

Table 3-3.1 Conipanent size and source contribution for a hypothetical fish

% Source contribution % Body hydrogen Components gl____....- in each Water Organic food conponen t

Tissue water 80 Tissue bound Exchangeabl e 100 0 6 11

Nonexchangeable GOa 40a 14b ._ ...... I-- ___ ._I__ aData of Patzer (1973).

%ata of Siri and Evers (1961). 93 with envi ronnlental tri ti uni levels. As a result, this nonexchangeable tritium component does not reflect current exposure levels unless

1 ong- term steady-state condi tion has been maintained under constant envi ronrnental levels.

3.3.4 Tritium Concentration in Aquatic Food Chains

3.3.4.1 Tissue-Water Tritium

Bioaccumulation fdctors reported for tissue-water tritium in aquatic organisms range from 0.58 to 1.1 with a mean value less than 1 (Bruner, 1973; Elwood, 1973; Harrison and Koranda, 1973; Stewart et a1 . , 1973; Weinberger, 1953; Weinberger and Porter, 1953). Thus, on the basis of thzie empirical studies, unity represents a conservative approximatiim of the tissue-water tritium bioaccumulation factor for aquatic organisms.

3.3.4.2 Tissue-Bound Tritium

Both laboratory and field measurements indicate that the bioaccumulation factor for the total tissue-bound component is less than 1. The tissue-bound t*-itium concentrations in plants and invertebrates have been summarized from sixteen references by Bruner (1973). In only two 0-T these cases were tissue-bound tritium concentrations higher than ambient water (Cohen and Kneip 1973; Koranda, 1965). These two measurements were made in areas of pulsed tritium inputs in the vicinity of nuclear power facili- ties or test sites and probably do not. represent steady-state 94 situations. Other field and laboratory measurements on tissue-bound tritium in plants and vertebrates showed tritium bioaccumulatian factors below unity (Elwood, 1973; Harrison and Koranda, 1973;

Kanamawa et ai., 1972; Porter, 1954; Rosenthal and Stewart, 1973; Stewart et a1 . , 1973; Strand et al. , 1973; Weinberger, 1953; Weinberger and Porter, 1953). Most studies did not include measurements of exchangeable triti um and nonexchangeable triti mi as separate components. Thus, on the basis of empirical data alone, it is not possible to determine a separate bioaccumulation factor for exchangeable tritium and nonexchangeable tritium, or to determine the degree to which the specific activities of certain organic molecules might vary araund the mean value for the total tissue- bound component. For this reason subsequent discussions focus on physical and biochemical aspects of the two ti ssue-bound components and their effects on the bioaccurnulation factor for exchangeable tritium and nonexchangeable tritium, as we71 as individual molecules.

3.3.4.2.1 Exchangeable Tissue-Bound Tritium

Movement of tritium into and out of the exchangeable component involves chemical exchange reactions as discussed in Section 3.3.2.

A variety of organic groups cornrnon in living tissue have been studied

to determine the steady-state concentration obta ned via exchange with

tritiated water (Weston, 1973). The results, wh ch are summarized in Table 3.3.2, indicate that tritium is not likely to pool in exchangeable 95

Table 3.3.2 Steady-state concentration in organic groups cornmon in living tissue

Organic group 1a Ref e ren ce

OH Groups

CH3COOH 1 .Q 3.4 CH30H 0.98 3.9 Cell ul ose 1.43 3.14 NH Groups

NH3 0.94 3.8 HCONH2 0.97 3.1 Ribonuclease 1 .Q 3.6 Dt-polyal anine 1.2 3.7

SH Groups

HZS 0.30 3.76 CH3SH 0.30 3.11 CH Groups

HC-CCH3 0.78 3.15 CH3SH 0.64 3.30

aa-1 ~ (Specific Activity organic group)/ (Specific Activity water). Since c1-I is based on specific activities it is independent of variations in hydrogen content, and must be multiplied by the percent hydrogen in the tissue and divided by the hydrogen content of water (11%) to produce a "dry weight" bioaccumulation factor for the organic group. components because in most cases the specific activity* of the molecules is less than that of the water.

A food chain is diagrammed in Figure 3.3.2 with organisnis compartmentalized as in Figure 3.3.1 and each consumer utilizing the organism below him as the food source. In both Figure 3.3.1 and 3.3.2 there is no pathway to accommodate the Fosd-chain transfer of exchangeable tri tiurn even though this obviously occurs. The effect, of food-chain transfer of exchangeabl E tritium is unimportant because 1) the exchangeable tritium of each organislil has ready

CIGC~SS to the same ambient tritium concentrations via exchange with tissue water, and 2) most aquatic organisms eat only a sillall fraction of their body weight per day thus the turnover of exchangeable tri ti urn appears rapid in compari son.

3.3.4.2.2 Nonexchangeable Tissue-Bound Tritium

At the base of the food chain in Figure 3.3.2, photosynthesis, a1 ong wi th other reduction reacti ans , incorporates tri ti urn from tissue water into the nonexchangeable component of plants. Data on the total tissue-bound cornponetit in plants indicate that there is a 3 to 55% discrimination against incorporatian of tritium rela- tiv? ta pratium (Bruner, 1973; Harrison and Karanda, 1973;

Rosenthal and Stewart, 1973; Weinberger, 1953; Weinberger and Porter, 1953). These deterini nati ons , however, did not differ- entiate between the two tissue-bound components.

"Specific activity is used here to man the ratis of radioactive to stable atoms of an element in a sample. 97

ORNL-0 WG 74-4497R2

CARNIVORE

AMBIENT WATER HERBIVORE 1 1 ti J

PLANT

TWT= TISSUE WATER TRITIUM ET = EXCHANGEABLE TRITIUM NET= NONEXCHANGEABLE TR7TIUM Figure 3.3.2 Three-1 ink food chain diagram with organisms compartmentalized, 98

Rambeck and Bassham (1973) and Kanaaawa et a1 - (1972) investigated the effects of oxidation and reduction reactions on specific acti vi ties of i ndi vi dual organi c molecules produced during the Krebs Cycle in algae. After five generations in a closed system the average speci fi c acti vi ty of the indi vi dual rnolecul es investigated was 30% less than the water to which they were exposed. Only two out of the 14 molecules investigated showed specific activity ratios greater than 1, citrate 1.8 and fuinarate 1.2. lhese results, although limited to 2 species of algae and the Krebs Cycle, indf- sate that riiost organic molecules discriminate against tritium incorporation -into nonexchangeable sites and that sigriifi cant increases in the specific activity of certain molecules due to the kinetics of nonexchangeable tri ti um are not 1 i kely . Further studies are needed to verify these results for various series of metabolic reactions in other organisms.

Consumer organisnis obtain nanexchangeable tritium by assimilation of intact tritiated food molecules as well as by reduction of tissue organics with tissue-water hydrogens. It can be calculated froiii Patzer's (1973) data. on herbivorous fish that the nonexchangeable component receives approximately 50% of its hydrogen From tissue-water hydrogen and 40% froiii food. Data are not available on the exact degree to which the specific activity of tile nonexchangeable component is affected by food-chain transfers of nonexchangeable tritium in food molecules, or by various oxidations and reductions. 99

Although empirical data on bioaccumulation factors for nonexchangeable tritium in organisms are wanting, it is possible to establish a maximum bioaccumulation factor for the nonexchangeable component by using the tissue-bound compartment sizes from Table 3.3.1, an average specific acti vity ratio for the exchangeable component from Table 3.3.2, and bioaccumulation factors determined empirically for the total tissue-bound coniponent in numerous freshwater organisms. Based on Table 3.3.2, 0.7 represwts the best estimate of an average bioaccumu? ation factor for the exchangeable component. Thus, the empirical ly-derived bioaccumulation factor of less than 1 for the total tissue-bound component, corsisting of 30% exchangeable tritium and 70% nonexchangeable tritium, requires that the bioaccurnulation factor of the nonexchangeable component be no greater than 1.1. 100

References for Section 3.3

The nuiiibers in the left-hand margin correspond to the reference numbers used in the tables of section 3.3.

3.1 Brokska'i , A. E. "The exchange of hydrogen with deuterium in solution," Trans. Faraday Sac. 33, 1180 (1937). 3.2 Bruner, H. D., Distribution of tritium between the hydrosphere and invertebrates, IN Tritium, A. A. Moghissi and M. W. Carter (eds.), pp. 303-313 (1973), Messinger Graphics, Phoenix, TlD CONK-71 0809. 3.3 Cohen, L, K., and P. J, Kneip, Environmental tritium studies at a PWR power plant, IN Tritium, A. A. Moghissi and M. W. Carter (eds.), pp. 639-646 (19731, Messinger Graphics, Phoenix, PID CQNF-710809 a 3.4 Eidinoff, M. L., et al., "Equilibrium between protium, deuterium, and tritium in the system: Hydrogen-acetic acid; isotopic fractionation factors in a catalytic hydrogenation ,'I J. Amer. Chem. Soc. 74, 5280 (1952).

3.5 Elwood, J. W., Tritium behavior in fish from a chronically contaminated lake, IN Radionuclides in ecosystems, D. J. Nelsaii (ed.) , pp, 435-439 (1973) , Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee, 1971.

3.6 Englander, S. W., "A hydrogen exchange method using tritium and sephadex: Its application to ribonuclease," Biochein. I2, 798 (1 963) * 3-7 Englander, S W., and A. Paul sen, Wydrogen-tri tiurn exchange of the random chain polypeptide, Biopolymers 7., 379-393 (1969). 3,8 Gutman, J. R., and 61. Wolfsbery, "Equilibrium in the exchange of tritium between ammonia and hydrogen and the zero-paint energy difference between NH3 and NH2T," J. Chem. Phys. 33, 1592 (1960).

3.9 Hallford, J. O., and E. Pecherer, "'Exchange of deuterium between methanol and water. Vibrations of the hydroxyl group in methanol and methanol-d. The entropy of methanal," J. Ckem. Phys. 6, 571 (1938).

3.10 Harrison, F. and J. J, Koranda, Tritiation of aquatic animals in an experimentalL, freshwater pool, IN Radionuclides in ecosystems, D. 3. Nelson (ed.), pp. 425-434 (1973), Proceedings of the Third Nati onal urn on Radi oecol ogy , Oak Ridge , Tennessee May 10-12, 1971.Symposi 101

References for Section 3.3 (continued)

3.71 Hobden, F. W. et a1 ., "'Determination and calculation of the equilibrium constants for isotopic exchange in the systems n- amyl alcohol -water and ethyl thiol -water. Vapor pressures and raman spectra of ?-amyl deuterio-alcohol and ethyl deuteriothiol p'' J. Chem. Soc. -61 (-1939). 3.12 Kanazawa, T., K. Kanazawa, and J. A. Bassham, Tritium incorporation in the metabolism of Chlorella pyrenoidosa, Environ. Sei.

Techno]. I6, 638-642 (1972).

3.?3 Koranda, J. J., Preliminary studies of the persistence of tritium and carbon'-14 in the Pacific Proving Grounds, Health

Phys. I11, 7445-1457 (1965).

3.14 Lang, A. G., and S. 6. Mason, Tritium exchange between cellulose and water:R. Accessibility measurements and effects of

cyclic drying, Can. 3. Chem. I38, 373-387 (1960).

3.15 Mahadevan, E. G., and E. V. Monse, "Equilibrium constant for the deuterium exchange between methylacetylene and water,'i 3. Chem. Phys. -43, 2556 (1965). 3.16 Marx, D. (Summarized by) "Etude des Constantes des Equil ibres Isotopiques du Deuterium entre 1'Eau et les Hydrures des Metalloides de la Deuxierne Famille," C.E.A. Report 1382, Sacl ay France 1960.

3.17 Patzer, R. e., A. A. Moghissi, and D. McNelis, Accumulation of tritium in various species of fish rearedN. in tritiated water, IN Proceedings of IAEA/NEA/WWO Symposium on Environ- men tal Behavior of Radi onuc? ides Re1 eased in the Nucl ear Industry, Aix-en-Provence, France, May 14-18, 1973, MEA/ SM-172/29 ., Vienna , 1973. 3,18 Pinson, E. A., and W. H. Langham, Physiology and toxicology of

tritium in man, J. Appl. Physiol. I10, 108-126 (1957).

3.19 Porter, J. W., Incorporation of hydrogen isotopes into various fractions of rapidly growing algae, I1v Biology Research Annual Report 1953, pp. 77-78, Hanford Atomic Products Operation, Richland, Washington HW-30437 (January 1954).

3.20 Rambeck, A., and A. Bassham, Tritium incorporation and retention in photosynthesizingJ. algae, Biochem. Biophys. Acta --304, 725-735 (1 973). 102

References for Section 3.3 (continued)

3.21 Raney, F., and Y. Vaadia, Movement and distribution of THO in

tissue water and vapor transported by shoots of HelianthusI and

Nicotiana,.-_1 Plant Phys. _I40, 383-388 (1965).

Rosenthal , Jr., and M. L. Stewart, Tritium incorporation in algae andG.. transfer M., in simple aquatic food chains, IN Radio- \-- nuclides in ecosystems, Ne1 son (ed, ) , pp. 440-444 (1973) Proceedings of the ThirdD. National J. Symposium on Radioecology, Oak Ridge, Tennessee, May 10-12, 1971.

3.23 Siri, W., and J. Evers, Tritium exchange in biological systems, IN IAEA Proceedings of the Symposium on the Detection and Use of Tritium in the Physical and Biological Sciences, Vol. 23 pp. 71-84 (1961).

3.24 Smith, T. E., and R. T. Taylor, Incorporation of tritium from tritiated water into carbohydrates, lipids, and nucleic acids, USAEC Report UCRL-50781 (NOVeinber 26, 1969). 3.25 Stewart, M. L,, G. M. Rosenthal, and J. R. Kline, Tritium discrimination and concentration in freshwater microcosms , IN Radionuclides in ecosystems, J. Nelson (ed.), pp. 452- 459 (1973), Proceedings of the ThirdD. National Symposium on Radioecology, Oak Ridge, Tennessee, May 11-12, 1971.

3.26 Strand, J. A., W. L. Templeton, and E. G. Tangen, Accumulation and retention of tritiurn (tritiated water) in embryonic and larval fish, and radiation effect, IN Radionuclides in ecosystems, D. J. Nelson (ed.), pp. 445-451 (1973), Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee, May 10-1 2 1971 . 3.23 Thompson, R. C., and 3. E, Ballou, Studies of metabolic turnover with tritium as a tracer. V. The predominantly nondynamic state of body constituents in the rat, J. Biol. Chem. --223 795-809 (1956).

3 -28 Weinberger , D. , Met.abo1 ism of hydrogen isotopes by a1 gae IN Biology Research, Annual Report 1952, pp. 72-74, Hanford Atomic Products Operation, Richland, Washington, (1 953) . HW-28636 3.29 Weinberger, D. , and J. W. Porter, Incorporation of tritium oxide into growing @hlore??apyrenoidosa cells, Science x,636-638 (1953). 3.30 Weston, R. E. Jr. , Kinetic and equilibrium isotope effects of tritium substitution, IN Tritium, A. A. Moghissi and 61. Carter (eds.), pp. 289-300 (1973) , Messinger Graphics, Phoenix,W.

TID-CONF-71O809 I) 3.4 IODINE

3.4.1 Environmental Iodine The stable iodine content of freshwater ranges from 0.4 to 10.0 ppb depending upon the water source and the environment through which the water travels. Streams originating from glaciers or runoff are low in iodine, less than 1 ppb, while those originating from springs and ground water sources are higher, up to 5 ppb. Livingston (1963) calculated a median river water concentration of 2 ppb. Table 3.4.1 lists some representative stable iodine concentrations from various freshwater, marine, and geologic sources. Iodine in oceans is approximately 50 ppb and apparently originates fron the erosion of land masses (Bowen, 1966). The dominant physiocochernical form of iodine and its relative biological availability in freshwater is uncertain (Winchester, 1970).

3.4.2 Iodine Metabolism The thyroid removes inorganic iodine from the blood and uses it in the production of thyroxine, a hormone necessary for maintaining a variety of metabolic and growth fuwtions. The thyroid releases thyroxine to the blood stream where it combines with large proteins to form protei n-bound odine. Thyroxine is spl t from the large protein and enters the body tissues where it is degraded. Inorganic iodine produced by the degradation of thyroxine re-enters the blood and is eliminated by urinary excretion. Some evidence indicates that an unspecified portion of this may be reabsorbed by the thyroid (Chavin and Bowman, 1965). 104

Table 3.4.1 Iodine concentrations in surface waters, the ocean, and geologic sources

Average Mater Range Sample Concentration Reference (?Pb) (PPb) -- -.-_- Geman lake 3.5 2.5 - 4.9 4.18 Canadi an 1a ke 1.154 0.49 - 2.9 4.21 lake Michi gan 1 .Q 4.10 -e-- 4.39 Lake Mi chi gan 5.0 Mi chi gan streams 1 or less ...I-- 4.39

Missouri spring 4.7 or less --^-- 4.12 New Jersey sp ring 5 -2 or less ---- 4.12 Swiss rivers 9.56 0.4 - 1.3 4.15 Panama rivers 1 or less 1.0 - 0.6 4.14 River Average 2 .cJ < 1.0 - 10 4.28 Oceans 50 -..-." 4.57 $ai1 s ...--... so0 - 6000 4.47 Rock

m--- IgnEtOUS 568 4.5 Limes tans 1200 -_-- 4.5 %ha1e 2200 .....-1 4.5 ...... -- Coir 1 6099 4.5 1Of;

Although there are numerous studies on the physiological role of iodine, few of these investigations provide data on elimination rates and turnover times for iodine in aquatic organisms. The few studies that were found are summarized in Table 3.4.2. The rapid component

(C-1) is taken to represent plasma clearance through urinary excretion, while the slower component (C-2) is attributed to the thyroid.

3.4.3 Review of Iodine Bioaccumulation Factors

3.4.3.1 Stable Iodine Bioaccumulation Factors for Fish Tissue

Only the thyroid tissue and ovaries (eggs) show great iodine bioaccumulation. No studies were found that provided data allowing direct calculation of bioaccumulation factors for the thyroid tissue because thyroid stable iodine measurements for freshwater fish were not expressed as concentrations. Data were available on thyroid iodine concentrations in some marine species and these marine data are cited to corroborate our indirectly estimated freshwater thyroid bioaccumulation factors. Muscle bioaccumulation factors of freshwater and marine fish are very similar despite the much higher iodin2 content of oceans (Tables 3.4.3 and 3.4.4). Organic and inorganic iodine show the same percent distribution in the serum of estuarine flounder and freshwater whitefish (Hickman, 1962). The similar iodine bioaccumulation factors for muscle and the similar iodine metabolism in fish from freshwater and marine habitats provide some justification for comparing bloaccumulation factors of marine and freshwater species. References on stable iodine in fish ovaries did not report com- parabJe iodine water concentrations, thus, ovary bioaccumul ation factors 106

mm Lnl -e-- mm NI .. .. -1 e* el

0 100 u3m W CnQ.. d

I1 II I1 II I1 I1 II II II I1 II I1 ino- II II It II II I. I II me I1 II II I1 II II I1 11 I1

QW ht c' .. .I ma.. NCO c- I me- mcc) F

cu N I l V 0 Table 3.4.3 Stable iodine in freshwater fish muscle

Muse1 e Water Stable Species Concentration Concentration Iodine Location Reference ( 118/ kg 1 hg/1 iter) B Fa Switzerland Salmo fario 36 36 4.16b 7Salm fario 48 4% Germany 4.4b Salmo fario 2$ 24 New Zeal and 4 a 20b Salrno 50 50 New Lealand * mb --I_ fario Sal vel inus namaycush 10 10 Lake Erie ;: Gb --Luciopimelodus pati 30 30 4.32b

PiInref odus a1 bidus 40 40 4.325 ~orr,is inci sor 40 40 Mississippi River 4.45b A9I-..- _I-.- Microperus. sal moi des 50 50 Potomac River 4.45b s anr,ui aris 10 10 Mi ss i ss ppi River -I__-Pornoxi i 4.45; Perca f 7 avescens 20 20 Potomac River 4.45

-...-.-I -...-.-I Perm fl uviati 1 is 20 20 Swi tzerl and 4.16b Frcsiwater species 40 40 4 * 33b in general Carp 17 17 4.6 Fjcropterus sctlmoi des 30 30 4.6 Ei\fveTi nus nainaycush 31 31 4.6 River perch 40 40 4.6 15 15 Bl ack Rfver a Michigan 4 39 40 40 Paclfle: Coast Stream 4,26 IOncorhynchus kisutch 120 120 Lake Michigan 4.11C Perm flavecce~~. 120 120 Lake Michi gan 4,W Sal vel i EUS nanaycush 170 170 Lake Mi chi gan 4,llC Salmo trutta 110 110 Lake Michigan 4.1 Coregonus cl upeaformis 140 140 Lake Michigan 4.11ClC _I Me an 50 -.t 45 aTissue concentration reported without comparable water values was divided by 1 pg/liter to approximate a coriservati ve bjoaccuniulation factor. bCi teci i’n Vinogradov (1953). ‘Tissue concentratiom represent averages calculated from Copeland’s raw data that clearly specified that the tissue analyzed was muscle. 108

co d

ln Y

cu e- h

e---

M B,

P0 u 7 09 could not be directly calculated frm the studies. However, we were able to indirectly estimate an average ovary bioaccumulation factor for stable iodine from studies which reported both ovary and muscle iodine contents in the fish collected. From these data an ovary:muscfe ratio was calculated and mu7 tip1ied times the average muscle bioaccumulation factor to estimate an average ovary bioaccumulation factor for stable iodine. Data on stable iodine in fish wuscle were adequate to determine a reliable bioaccumulation factor. Due to the limited data on iodine in fish thyroid and ovaries, and the manner in which these data were reported, our estimates for thyroid and ovary bioaccumulation factors represent approximations.

3.4.3.1.1 Stable Iodine Bioaccumulation Factors for Fish Muscle Stable iodine bioaccumulation factors for fish muscle from various species of freshwater fish are listed in Table 3.4.3. BiQa~c~mulati~~

factors from a variety of investigators varied from 10 to 50 while determinations by Copeland et al. (1973) ranged from 120 to 170. By

using a 5 ppb Lake Michigan iodine concentration, as reported by

Robertson and Chaney (1953), with Copeland's tissue iodine concentrations, a range of bioaccumulation factors can be calculated for Copeland's samples that agrees well with ranges reported by other investigators.

However, Winchester (14701, reported 1 ppb or less as the iodine concentration in Lake Michigan, thus, indicating that Copeland's water concentrations are correct and that higier tissue iodine concentration and higher bioaccumulation factors for Lake Michigan fish cannot be explained by an underestimate of iodine concentration in water. At this time it is the opinion of the authors that although Copeland's 110 data are higher than the others, we must accept their bioaccumulation factors of 120 to 170 for stable iodine in .freshwater fish muscle until other data clarify the issue. From data in Table 3.4.3 we calculate an average bioaccumulation factor of 52 for stable iodine in fish muscle with a range from 10 to 170. Based on clear decreases in iodine tissue concentrations from marine to anadromous to freshwatev- species, it does not appear that homeostatic control is a factor in controlling muscle iodine levels in .fish (Table 3.4.4). Anadromous fish sampled during their spawning run show stable iodine tissue concentrations higher than strictly freshwater species (Table 3.4.4). This is evidently due to their recent exposure to higher iodine concentrations in oceans. Fontain et a1 . (1948) , found 1280 pg/g of iodine in the serum of At1antic salmon at the beginning of their migration spawn. At the end of migration, serum iodine had dropped to 280 ug/g. It was not determined if iodine in other tissues showed a similar decline. The drop in serum iadine could be due to lower freshwater iodine concentrations and a lack of homeostatic control, or to a combination of physiological factors.

3.4.3.1.2 Stable Iodine Bioaccumulation Factors for Fish Thyroids and Ovaries In rainbow trout and most teleosts the thyroid is not assembled into a single gland but, instead, is scattered throughout the lower jaw. In some cases thyroid tissue may occur in the eyes and kidneys as well (Baker et al., 1955). The removal of diffuse thyroidal tissue without portions of nonthyroidal tissue is difficult. As a result, 171 iodine concentrations are either not expressed on a unit weight basis or qualified to the extent that the tissue may contain nonthyroid material (Short et al., 1969). Robertson and Chaney (1953) an'alyzed sections of careful ly dis- sected Lake Michigan rainbow trout tissue excised from the floor of the mouth from the first to the fourth gill arch. The iodine content was expressed as total micrograms of iodine in the total "thyroid mass" (Table 3.4.5.). The weights of these samples were not reported.

No determinations were made of the stable thyroidal iodine content, expressed per unit of exclusively thyroid tissue of freshwater fish.

Using 0.003 g/100 g of fish as an approximation of the mass of the rainbow trout thyroid (Spector, 1956), we calculated an approximate thyroid iodine concentration of 269 yg/g by using the average weights of the trout used in Robertson and Chaney's (1953) study. Dividing this thyroid concentration by Copeland's Lake Michigan iodine concentration of 1 ppb gives an estimat2 of 270,000 for the trout thyroid bioaccumulation factor. To provide some verification of this thyroid bioaccumulation factor, an average thyroid bioaccumulation factor for marine fish of 770,000 was calculated using data compiled from Vinogradov (1953) in Table 3.4.6. Considerin: the approximate nature of our estimates, these two values are fairly close and tend to support the freshwater thyroid bioaccumulation factclr. Hickman (1962) measured inorganic iodine in the serum and thyroids of estuarine flounder and found that the serum levels of iodine increased on a seasonal cycle as the salinity of the water increased. However, in the case of thyroidal iodine, only inorganic iodine showed a clear 112

Table 3.4.5 Total iodine present in the thyroid and eggs of five rainbow and Cali- fornia steelhead trout (in

Rainbow Trout Steel head Troutb

-I ^- Thyroid Eggs Thyroid Eggs

3.5 38.6 367 1190 8.5 18.6 755 2323 12.3 41.3 377 1814

11.9 70.0 --I 7 749 l6,Q 55.0 - -- 106%

10J 44. oc sod 1628 aRobertsan and Chaney (1953). bStee7iliead trout are rainitpo anadromousThey were trout (Salrno gardneri). se- cured from the mouths of Ca? ifornia rivers to up 150 miles upstream,

‘Mean values (I 113

Table 3.4.6 Thyroid stable iodine bioaccumulation factors for marine fish

Species ThyroiI odi ne Thyroi d Referenceb BF~ (PPd

Gadus 2,800 56,000 4.9,4.30 aeslefinus

Gadus 11,600 232,000 4.9,4.31 mor rhua

Salmo eriox 173,600 3,500,000 4.30,4.31 Salmo salar 9,600 192,000 4,9,4.31 S combe r 32,200 644,000 4.29,4,3C) scombrus

Sebas tes 46,200 924,000 4,29,4.31 mari nus

P1 euronectes 7 4,9,4,31 81atessa 7,800 56,000

23,800 476,000 4.9 uvu1 saris - Mean 772,500

aCalculated using 50 ppb as marine water iodine concentration (Table 3,4.1). bCi ted in Vinogradov (1953). 114 response to environmental iodine fluctuations. The protein-bound iodine, which constitutes 95% of thyroidal iodine, varied much less and with no apparent pattern. Protein-bound iodine was, in fact, highest in September when environmental iodine was at a law point.

Thyroid iodine concentration appears to fluctuate 1 i ttle in response to environmental iodine. Whether this is due to homeostatic control, slower turnover rates, or other factors is uncertain. From Table 3.4.5 it is apparent that the ovaries (eggs) contain most of the iodine present in female rainbow trout. Due to the manner in which the data were reported and to uncertainties concerning analytical procedures, we calculated an ovary:muscle concentration ratio for stable iodine (Table 3.4.7) and then multiplied this ratio by the average muscle bioaccumulation factor from lable 3.4.3 to estimate a stable iodine bioaccumulation factor of 800 for ovaries.

3.4.3.2 Iodine-131 Bioaccumulation Factors far Fish Tissue

3.4.3.2 a 1 Iodine-131 Bioiaccumul ation Factors for Fish Fluscl e

The few experiments reporting iodine-131 levels in fish muscle are listed in Table 3.4.8. Experiments by Short et al. (1969) indicate that the food web is the principal mode of uptake. One study by Hunri and Reineke (1964) indicates that iodine can be accumulated directly from the water. In Short's study, an aquarium and a lake were both tagged with iodine-131. After 29 days fish in the lake showed a muscle bioaccumulation factor of 100, while those in the aquaria, without their natural food sources showed a bioaccumulation

.factor of 1.7. It is possible, although experimental procedures were 115

Table 3.4.7 Stable iodine bioaccumulation factor for fish ovaries (eggs) derived from relative concentrations of stable iodine in muscle and ovaries

Ovar.y/rnuscl e Ovary a Species R'atio BF Reference

Salrno gardneri 15 750 4.39

Sal mo gardneri 14 700 4.39 Salm fario 29 1450 4.16 Perca fluviatalis 6 -300 4.16 Mean Ovary BF 800 a Ovary bioaccumulation factor = Ovary/muscle ratio times

average muscl e bioaccurnul ation factor from Tab1 e 3.4.3 a 116

Table 3.4.8 lodine-131 bioaccumulation factors fur muscle in freshwater animals

Iodine-l 31 Species aF Location Reference -- Carp 25 Lake 4.27 Cyprinus carassi us mole fish) Prochi 1odus 0.5 Aquaria 4.3 Pime1 odus 0.8 Aquaria 4.3 Ci chl asoma 0.5 Aquaria 4.3 Miscel 1aneous 20 hean) Aquaria 4.43 aquatic 17 species 1eech gastropods crustaceans insect 1 arvae carp tadpo 1es

Carp 12 Aquaria 4.44 Cypri nus carpi o Salrno qai rdneri 100 Lake 4.40

Mean Fish Muscle BY ..-._I %See Sec. 3.4.3.2.1. 117 not described, that the low bioaccumulation factors reported by Beninson (1966) for fish in aquaria studies were due to the absence of a contaminated food source. The low bioaccumulation factors do not appear reliable in view of Short’s study and the much higher stable iodine bioaccumulation factors reported in Table 3.4.8. An average iodine-131 bioaccumulation factor of 39 for fish muscle is calculated for fish muscle from Table 3.4.8 (Beninson’s data omitted). Because of the 8-day’ radioactive half-life of iodine-131, the turnover rate of iodine in a spec-ific tissue and the length of the primary contamination pathway control the degree to which bioaccumulation factors for iodine-131 are less than bioaccumulation factors for stable iodine. Iodine-131 absorbed directly from the water wjll have less time to decay than iodine-131 obtained from food. Section 2.4 and Equation (2.4.9) describe a method for calculating bioaccumulation factors for radioactive isotopes based on their stable element bioaccumulation factors. Using Equation (2.4.9), the average stable element bioaccumulation factor for fish muscle (Table 3.4.3), and an average fast component elimination coefficient of 0.43 (Table 3.4.21, we calculate an iodine-131 bioaccumulation factor of 43. Based on the similarity between this calculated bioaccumulation factor and the average bioaccumulation factor determined from empirical studies, we recommend an average bioaccumulation factor of 40 for iodine-131 in fish muscle.

3.4.3.2.2 Iodine-131 Bioaccumulation Factors for Fish Thyroids Short et al. (1969) removed the lower terminus of the gill arch, including the dorsal aorta and surrounding muscle, from lake trout 118 removed from an iodine-131 tagged lake after 29 days of exposure. An iodine-131 bioaccumulation factor of 12,000 was calculated for these "thyroid" samples which contained unspecified amounts of other tissue.

This value appears low when compared to the thyroid bioaccumulation factor of 270,000 for stable iodine (Sec. 3.4.3.1.2). Since no freshwater iodine-131 bioaccumulation factors have been determined for thyroids free from attached nonthyroidal tissue, we recomnend an iodine-131 thyroid bioaccumulation factor of 110,000. An ll-day elimination coefficient for the long component in trout (Table 3.4.2) and an average stable iodine bioaccumulation factor for thyroid tissue of 270,000 were used in calcul ating the recommended thyroi d bi oaccuinul ation factor . Further research is needed to clarify this issue.

3.4.3.3 Stable Iodine Bioaccumulation Factors for- Plants The only stable iodine determination found for freshwater plants was an average bioaccumulation factor of 800, calculated by Copeland and Ayers (1970) for samples of phytoplankton (green and bluegreen algae) taken from Lake Michigan.

3.4.3.4 Iodine-131 Bioaccumulation Factors for Plants Bioaccumulation factors for rooted and floating macrophytes are combined in Table 3.4.9 since they are within the same range. Iodine- 131 bioaccumulation factors for freshwater algae from one study awr- aged 255 (Short et a1 . , 1969).

3.4.3.5 Stable Iodine Bioaccumulation Factors far Invertebrates Table 3.4.10 lists stable iodine bioaccumulation factors for a small number of aquatic invertebrates. In insects, and probably Table 3.4.9 Iodine-731 bioaccumulation factors for freshwater plants

Species Remarks Reference BF Macrophytes 130 Tarik 4.3 178 Tark 4,3 145 Tank 4.3 'I 78 Tank 4.3 162 Tank 95 Tar k 100 Tarik 289 Tank 7: ?a:?k 4.44 154 Tank a 44 26 Tak .44 -is 134 TX&. 4.44 60 Talk 165 Tank Carex P 101 Tan ph aqnum k S 90 Lake Nupkar 60 Lake Mean of 32 93 Tank Plant Spec ies Mean Macrophyte BF 199

A1 gae 388 Lake 4.48 SDi roqyra and 13 Lat

e 0 Q 0 en P pl I--- r-- w d e 9$ d .rt-

sa, I I I I I F-- M I I I I I G-- M I I I I I .C I I a I I I r ti- v3

0 B 0 6 7 m Qsa 0 N m Wm 0 T-

n I I uf, I 0 I

W I m m.-. 0 uLn 0 I Ln I $-" usLn mBI I

A crustacea as well, most of the iodine appears to be concentrated in the integuement. Iodine-131 studies indicate that little iodine is lost from insects except when the integuement is discarded (Odurn and

Golley, 1963; Van Hook and Crossley, 1969) a The average bioaccumulation factor for crustacea, the only group of invertebrates for which stable iodine values of soft tissues or whole animals were found, is 343.

3.4.3.6 Iodine-131 Bioaccumulation Factors for Invertebrates Table 3.4.1 1 lists iodine-131 bioaccuniulation factors for various invertebrates. An average bioaccumulation factor value is recorded for aquatic insects and molluscs. The difference in the lake bioaccumulation factor of 140 and the aquaria bioaccurnulation factor of

600 reported for the amphipod is atti-ibuted to incomplete mixing of lake waters (Short et al., 1969). 122

Tab1 e 3.4.11 Iodine-131 bioaccumulation factors for invertebrates ~- Organism Iodine-131 Remarks BF Tissue Reference

Crustacea : Crayfish 10 Muscle Lake 4.40 (Paci fastacusj 62 Soft parts Lake 4.40 240 Carapace Lake 4,40 60a Whole Tank 4.40 140 Whole take 4.40 Aeql a Tank 4.3 Littoral Dlank tan 530 Lake 4.40 Liinneti c b1anktun 500 Lake 4.40 Aquatic Insects: Dragon fly larvae 151 ? Tank 4.48 (keucorrhinia) May fly larvae 690 ? Tank 4.45 (c1oeon) Caddis Fly larvae 163 ? Tank 4.48 (Glyphotael ius) Drone fly 'laiTv-3e 600 ? Tank 4.48 (EriI_-._____ stalis) I-_ Mean BF for 400 Aquati c Insects worm : Hespsbdel la 10 ? 4.38,4.44

Sponge : __Spncmi 11a 200 ? Lake 4.27

Mol 1us cs : Diplodon 11 ? Tank 4.3 @-35ia 23 ? Pan k 4.3 Planorbis 134 ? Tank 4.3

Mar9a- ri ti fera I_- 10 Mus cl e Tank 4.40 53 Soft Qarts Lake 4.40 Li mnaea 23 ? Tank 4 I 38 94.44 1% 4.38,4.44 -._IIWadi x ? Tank Ani sus 53 ? Tank 4. ?8,4.44 mbi__.-..._I____s 78 ? Tank 4.3%,4.44 ._._Drei ssensi a 140 Gills Tank 4.19 70 Marntl e Tank 4.19 40 Viscera Tank 4.19 20 Byssus Tank 4.19 400 She1 1 Yank 4.19 Pond sndil 20 ? 'Tank 4.43 Mean BF for 50 Mol 1 uscs 123

References for Section 3.4

The numbers in the left-hand margin correspond to the reference numbers used in the tables of section1 3.4.

4.1 Baker, K. F., 0. Berg, A. Gorbman, R. F. Nigrelli, and M. Gordon, Functional thyroid tumors in the kidneys of platyfish, Cancer Res. 118-123 (1955).

15, \ 4.2 Baptist, J. P., D. E. Hoss, and C. W. Lewis, Retention of Cr-51, Fe-59, Co-60, Zn-65, Sr-85, Nb-35, In-ml41, and 1-131 by the Atlantic croaker (Micropogon -undlulatus), Health Phys. -18, 147- 148 (1970).

4.3 Beninson, D. E., E. Vander Elst, and D. Cancio, Biological aspects in the disposal of fission products into surface waters, IN Disposal of radioactive wastes into seas, oceans, and surface waters, pp. 337-354, Proceedings of International Atomic Energy Agency Symposium, Vienna, 1966, STI/PUB/126. 4.4 Bleyer, B., Zur Kenntnis der Iodes als Biogenes Element, Biochem. Z. -170, 256 (1926). 4.5 Bowen, H. J. M., Trace elements in biochemistry, Academic Press Inc., New York, 1966. 4.6 Causeret, J., Fish as a source of mineral nutrition, IN Fish food, 6. Borgstrom (ed.), pp. 2C5-234, Academic Press, New York,as 1962.

4.7 Chavin, Thyroid distribution and function in the goldfish, CarassiusW. auratus L., J. Exp. Zool. 133, 259-280 (7956). _I

4.8 Chavin, and 8. Bouwman, Metabolism of iodine and thyroid hormone W.,synthesis in the goldfish, Carassius auratus L., (;en. comp. Endocr. -5, 493-503 (1965).

4.9 Class, K., Uber das Vorkommen des Jods im Meere und in Meeresor-

ganismen, Arch. Math. Naturv. I_40, 1-150 (1931) , If4 Vinogradav 1953.

4.10 Copeland, R. A., and J. C. Ayers, Trace element distribution in water, sediment, phytoplankton, zooplankton, and henthos of Lake Michigan: A base1 ine study with calculations of conccntra- tion factors and buildup of radioisotopes in the food web, Environmental Research Group Special Report No. 1, Great Lakes Research Division, Univ. Michiaan, Ann Arbor. 1970. 124

References for Section 3.4 (continued)

4.11 Copeland, R. A., R. H. Beethe, and W. W. Prater, Trace element distribution in Lake Michigan fish: A baseline study with calculations of concentration factors and equilibrium radioisotopes distribution , Environmental Research Group Special Report No. 2, Great Lakes Research Division, Univ. Michiydn, Ann Arbor, March, 1973.

4.12 Dundee, W., and A. Gorbman, Utilization of radioiodine by the thyroid of neotenic salamander , Eur cea tynerensi s- Moore- and Hughes, Physiol. Zool. 33, 58-63 T-y71960 . 4.13 Eales, J. G., The influence of temperature on thyroid histology and radioiodine metabol isin of yearling steelhead trout, Salmo

gairdneri.--_- , Can. J e Zool . I_42 829-841 (1 954). 4.14 Ewing, R. A., J. E, Wows, and R. B. Pri'ce, Bioenvironmental and radiological-safety feasibility studies Atlantic-Pacific Interoceanic Canal: Radionuclide and stable-element analysis of environmental samples from Routes 17 and 25, Battelle Memorial Institute, BMI-171-29 (1969) , also IOCS-MEMO-BMI-33 (1969). 4.15 Fell enberg, Th. von , Das Vorkorianen der- Kresl auf und der Staffwechsel des Jods, Ergebn. Physiol . _I__25, 176-363 (1926). 4.16 Fellenherg, Th. von, Untersuchungen uber dins Vorkorruieri v~nJod in

der Nature, Mitt, Lebensm. Hyg., Bern II_14, 161 (1923).

4.1 7 Fontain, M. , F. Lachiver, S, Leloup, and M. 01 ivereau La Fonction thyraidienne du Saurnon (Salmo salar L.) au cours de sa migration reproductrice, J. Physio-rm, - 182-84 (1948.). 4.18 Glasser, R., Zur Frage der Anreicherung von Radiojod (5-?31) in der Muschel Dreissensia polymorpha Pall., Naturwissenschaften 47, 284 (1960).

4.19 Glasser, R., Uber die Aufnahm van Radiojod aus wassriger Losung durch die Muschel Dreissensia polymorpha Pall., Zool. Jb. Physiol. -69, 339-378 (1961). In Polikarpov 1966.

4.20 Hercus, C. E., and K. C. Roberts, The iodine content of foods, manures, and animal products in relation to the prophylaxis of endemic goitre in Nw Zealand, J. Hyg. , Camb. 26, 49 (1927). 4.21 Hickman, C, P., Influence of environment on the metabolism of iodine in fish, Ann. N.Y. Acad. Sci., Gen. comp. Endow. Sup. 1, 48-62 (1 962) . 125

References for Section 3.4 (continued)

4.22 Hunn, J. B., and P. 0. Fromm, Uptake, turnover and excretion of iodine-131 by rainbow trout (Saimo gairdneri) , Biol. Bull. -126, 282-290 (1 964).

4.23 Hunn, J. B., and P. 0. Fromm, In vivo uptake of radioiodine by rainbow trout, J. Water Poll. Contr. Assoc. -38, 1981-1985 (1966), 4.24 Hunn, J. B., and E. P. Reineke, Influence of iodine intake on iodine distribution in trout, Proc. SOC. exp. Biol. Med. 115,-- 91 -93 (1 964).

4.25 Hunt, E. L., and J.? N. Dent, Iodine uptake and turnover in the frog tadpole, Physiol Zoo1 . -30, 87-91 (1957). 4.26 Jarvis, N. D., R. W. Clough, and E. 0. Clark, Iodine content of

Pacific Coast salmon, Univ. Washington Pub. Fish. I1, 109-140 (1925-28). In Robertson and Chaey 1953. 4.27 Kolehmainen, S., S. Takatalo, and J. K. Miettinen, A tracer experiment with iodine-131 in an oligotrophic lake, IN Radionuclides in ecosystems, D. J. Nelson (ed.), pp. 278-284 (1973), Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee, May 10-1 2, 1971 . 4.28 Livingston, D. A., Chemical composition of rivers and lakes, Geological Survey Professional Paper 440-6 (1963). 4.29 Lunde, Uber die Geochemie und Biochemie des Jods mit. besonderer BerucksichtigungG., der norwegischen Kropfprophy laxe, Wien. klin. Wschr. -40, 1559 (1928). In Vinogradov 1953.

4.30 Lunde, G., J. Boe, and K. Closs, Iodine content of American marine animals, J. Cons. Perm. int. Explor. Mor. -5, 216 (1930). In Vinogradov 1953.

4.31 Lunde, G., K. Closs, H. Haaland, and S. 0. Madsen, Jodinnholdet i

norsk fisk og fiskerprodukter, Aarsberetn, Norg. Fisk., No. I4, 1-41 (1928). In Vinogradov 1953.

4.32 Mazzocco, P., Iodine in food in salta, Sem. Med., B. Aires I37, 364 (1930).

4.33 Monier-Williams, G. W., Trace elements in food, John Wiley and Sons, New Work, 1950.

4.34 Nilson, H. W., and E. J. Coulson, The mineral content of the edible portions of some American fisheries products, Invest. Rep. U.S. Bur. Fish. -41 , 1-7 (1939). In VinDgradov 1953. 126

References for Section 3.4 (continued)

4.35 Odum, E. P., and F. B. Golley, Radioactive tracers as an aid to the measurement of energy flow at the population level in nature, IN Radioecology, V. Schultz and A. Klement (eds. ) , p. 403 (1963) , Proceedings of the National SymposiumW. on Radioecology, Fort Collins, Colorado , September 10-1 5 1961 . 4.36 Ohmomo, Y., H. Suzuki, M. Sumiya, and Saiki, Studies on accumulation excretion and distribution of iodine-131M. and cadmium-115m in freshwater fish and rnarine fish, J. Rad. Res. 13, 3 (1972). 4.37 Orr, J. El., and I. I-eitch, Iodine in nutrition, A review of existing information up to 1927, llnited Kingdom Medical Research Council, Special Report Series No. 123, His Majesty's Stationary Office, London (1929) .

G. of Aquatic____ Orggnisms (Reinhold 4.38 Polikarpov, G. , Radioecolo Book Div., N.Y., 19-

4.39 Robertson, 0. H., and A. L. Chaney, Thyroid hyperplasia and tissine iodine content in spawning rainbow trout: A comparative study of Lake Michigan and California sea-run trout, Physiol . Zool. -26, 328-340 (1953).

4.40 Short, R. F. Palurnbo, P. Olson, and J. R. Donaldson, The uptake ofZ. F.,iodine-131 by the biotaK. of Fern Lake, Washington, in a laboratory and a field experiment, Ecology -50, 979-989 (1969).

4.41 Spector, W. S., Handhook of Biological Data, W. B. Saunders Co., Phi:adelphia (1956).

4.42 Takatalo, S., S. Kolehtnainen, and J. K. Miettinen, Excretion of iodine from perch (Perca--._I.^ fluviati1 is), IN Biological enrichment chains of radionuclides in Finnish lakes, pp. 104-111 (1968). Annual Report for the Period Sept. 1, 1967 to August 31, 1968, Department of Radiochenlistry, Univ. Helsinki , Finland.

4.43 Timofeev-Resovskii, N. V., E, A. Timofeeva-Resovskaya, G. A. Milyutina, and A. B, Getsova, Coefficients of accumulation of radioisotopes of sixteen different elements by freshwater organisms and the effect of EDTA on some of them, Dokl. Akad. Nauk SSSR 132, 369-372 (1960). 4.44 Timofeyeva-Resovskaya, Ye. A., Distribution of radioisotopes in the main components of freshwater bodies (Monograph) , Tr. Inst. Biol. Akad. Nauk SSSR, Ural'skiy filia 30, 1-78 (JPRS 21.816)

1963 e 127

References for Section 3.4 (continued)

4.45 Tressler, D. K., and A. W. Wells, Iodine content of sea foods, Doc. U.S. Bur. Fish., No. 967 (1924).

4.46 Van Hook, R. I., Jr., and D. A. Crossley, Jr., Assimilation and biological turnover of cesium-134, iodine-131, and chromium-

51 in brown crickets Acheta-I -domesticus (L.), Health Phys. 463-467 (1969) . 16, 4.47 Vinogradov, A. P., The elementary chemical composition of marine organisms, Sears Foundation f'x- Marine Research, Yale Univ., New Haven, 1953.

4.48 Volkova, A., The build-up factor of the radioisotopes of some chemical elementsG. in aquatic insects, Entomologicheskoye Obozrenity Vol . 17, 516-519 (1963), Trans. 63-41234 (November 25, 1963). 4.49 Mil ke-Dorfurt, E. Uber den J3dqehalf der Schalen von Muscheltieren. Sum Chemischen Teil des Kropfaroblems. 11, Biochem. Zeit. -192, 74-82 (1 928).

4.50 Willis, L., and B. 8. Valett, Metabolism of iodine-131 in two amphibianD. species (Taricha- granulosa and Rana pipiens), IN Radionuclides in ecosystems, D. J. Nelson led.), pp. 394-400 (1973), Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennessee, May 10-12, 1971.

4.51 Winchester, J. W., Chemical equilibria of iodine in Great Lakes and waters, IN Proceedings of the Thirteenth Conference on Great Lakes Research, pp. 137-140, 1970, International Association for Great Lakes Research.

3.5 MANGANESE

3.5.1 Environmental Manganese

Manganese is found in six major forms in natural waters (Gibbs, 1973; Wangersky, 1963). In the suspended phase, manganese may be (1) a part of the crystalline structure of suspended material, (2) precipitated onto the suspended material probably as Mn02, (3) adsorbed onto solids by ion exchange, and (4) incorporated into solid biological materials. In the dissolved phase manganese may occur as one or more (5) dissolved ionic species, e.g., Mn ++ or inorganic associations, and as (6) complexes with organic molecules in solution.

Gibbs (1973) found that 10% of the manganese in the Yukon River and 17% of the manganese in the Amazon River were in the dissolved phase; less than 1% of the manganese in either river was on ion exchange sites.

In the Yukon River 46% of the manganese was precipitated as metallic coatings, 37% was in the crystalline structure and 7% was in biological material. In the Amazon River 50% of the manganese was on the precipitated metallic coatings, 27% was in the crys ta 1 1 ne structure and 5% was in biological material. In surface waters of five Alaskan akes and Par Pond (South Carolina),

51 to 94% of the manganese was in the p rticulat ~ phase (Marshal 1 and LeRoy, 1973). In lakes, most of the manganese in the particulate phase is probably MnOZ precipitates (Marshall and LeRoy, 1973). Dissolved manganese may be largely bound to organics since Mn f+ will most likely oxidize to Mn02 on surfaces under aerobic conditions (Gibbs, 1973;

Wangersky, 1963). Physicochemical form of manganese affects biological availability (Gibbs, 1973): (1) manganese in the crystalline structure 130

of sediments is not available .far uptake; (2) manganese in organic material s and manganese precipitates may be sornewhat available; and (3) manganese in solution or at in exchange sites may be very available.

Since? the proportions of these different physicochemical forms may vary considerably from site to site, the availability of manganese to organisms would be expected 'to vary from site to site. This picture is

further complicated by strong seasonal cycles of manganese concentra-

tion in these forms (Marshall and LeRoy, 1973)-

3.5.2 Manganese Metabo 1i sm

It was noted in Section 2-3 that the availability of a homea- statical ly-controlled element in different physicochemical forms had no effect on the concentration of that element in the organism. The binaccumulation factor of a homesstatical ly-control led stable element and the bioaccumulation factor of the radionuclide are given by Equa- tion (2-1.4). Fortunately, the considerable site to site variability in manganese availability is not an issue for many organisms since manganese is homeostatically controlled in vertebrates and in some invertebrates In a literature review Schroeder, Balassa, and Tipton

(1966) concluded that : "An efficient homeostatic mechanism for manganese appears to operate in a1 1 vertebrate and most invertebrate animals." Bryan and Ward (1965) suggested that the marine lobster

Homerus.. __. .- has a regulatory mechanism for manganese and further suggested that a similar mechanism should exist in niolluscs and fishes. Cavallerca and Merlini (1967) found only small site to site variations in manganese concentrations among aquatic vertebrates. Bortol i et a1 . (1969) found no differences in manganese concentrations among fishes from four different Italian lakes and Lentsch et al. (1973) detected little or no variation in manganese concentrations in fishes from the Hudson River, despite great changes in manganese concentration in the water.

3.5.3 Review of Manganese Bioaccu;nulation Factors

3.5.3.1 Manganese Bioaccumulation Factors for Fishes

As indicated in Section 3.5.2 the evidence for homeostatic control of manganese concentration in fishes is strong. Manganese concentrations in fishes and water and the bioaccumulation factors for manganese in fishes are tabulated in Table 3.5.1. Table 3.5.2 gives the means and coefficients of variation for the data on manganese concentrations in water and fishes given in Table 3.5.1. The small coefficients of variation for fishes in Table 3.5.2 and the large range of bioaccumulation factors in Table 3.5.1 support the hypothesis of homeostatic control of manganese. It is also interesting to note that manganese concentrations in estuarine and marine fishes are about the same as those in freshwater fishes. The average muscle concentration of manganese in Pomatomus saltatrix (bluefish) a marine epipelagic carnivore, is 0.22 ppm wet weight and the average concentration of manganese in Antimora

rostrata,I a bathy-demersal fish, is 0.21 ppm (Cross et al., 1973). The concentration of manganese in the ocean is about 0.002 ppm. Man- ganese concentrations in five species of estuarine fishes was 4.7 ppm on a whole fish basis (Cross and E?rooks, 1973). These values for marine and estuarine fishes fall within one standard deviation of the mean for freshwater fishes (Table 3.5.2). Table 3.5.1 Bioaccumulation factors for manganese in freshwater fishes

Stable element Stable element Species j-i ss ue ~~~dfng habits concentration concentration BF Location Reference in issue in water FPPm) (PPb) Five irrigated sitesb Alburnus alboreila whole plankton 20.6+0.7( 6) 9.0- 1 ooa 2300-21 0 5.4 {bleak) feeder Five irrigated sitesb iepoinis gibbosus who ie 15.7~0.9(5) 9.0- looa 2000-980 5.4 klburnus alborella whole plankton 12.7+0.8( 11 ) 9.0-100a 4 400-1 30 Five irrigated sites' 5.4 feeder Lepomi s gi bbos us whole 8.1+0.3(21) 9.0-10ga 900-80 Five irrigated sitesC 5.4 Zobi tis taenia who 1 e bot tom 5.220.4 (4) 9.0-100a 580-50 Five irrigated sites' 5.4 feeder whole bottom 4.89.7 (4) 9.0-1 0oa 530-50 Five irrigated sites' 5.4 feeder Tinca tinca whole scavenger 3.9L0.3 (6 1 9.0-1 0oa 430-40 Five irrigated sitesc 5.4 (tench) Average of Perca whole 7.6 4Za 180 Lake Varese, Italy 5.2 f 1 uv iat i 1 is-ch j , Scardinius erythro- whole 8.6 2 3a 374 Lake Camabbio 5.2 phthalnius (rudd), who 1 e 3.6 7 7a 27 'I Lake Maggiore 5.2 and Le onis CJ~ bbosus-E, whole 5.2 Sa 650 Lake Monate 5.2

Alosa pseudo- whole zooplankton 4.69.3 (19) 0.92+0.08( 53)6 5000 Lake Ml'chigan 5.6 feeder

WE) flesh 0.88+0.11(6)- 0.92+0.08(53f 960 Lake Michigan 5.6 reqontis flesh plankton and 0.25+0.03- 1722 (4)a 15 Lake Maggiore 5.17 m;;;:;;;;Co benthic feeder

Eondei 1 a)1""" bone 25 1752 (4)a 1500 Lake Maggiore 5.1 7

Alburnus alborella flesh plankton - 17+2 [4)a 21 Lake Maggiore 5.17 (bleak) feeder 0.36+0.04 bone 41 '1722 [4)a 2430 Lake Maggiore 5.17 Table 3.5.1 (continued)

-~ Stable element Stable element Tissue Feeding habits concentration concentration BF Location Species in tissue in water Reference ( PPm) (PPb) Scardi ni us flesh omni VO~OUS 0.3550.4 17+2- (4)a 21 Lake Maggiore 5.17 ery-.!w;yh tha 1mus bone 39 1722 (4)a 2300 Lake Maggi ore 5.17 unknown Notropi s hudsoni us whole 3.3+0,3( 15) 0.9250.08 (53) 3600 Lake Mi chi gan 5.6 (spottai 1 shiner) flesh benthic 0.57t- 0 * 06 ( 3) 0.9250.08 ( 5 3 1 550 Lake Michigan 5.6 flesh benthic 0,34+0.04 (5) 0.9220.08 (53) 370 Lake Mi chi gan 5.6

Core onus artedii flesh toopl ankton and 0,2920.05 (6) 0.9250,08( 53) 31 5 Lake Michigan 5.6 benthic + feeder -I w 'wii;j flesh 0,520.1 (3) 0.92~0.08(53) 500 Lake Michi gan 5.6 w P r~sobi urn flesh benthic 0.2520.03 (13) 0.92tO .08( 53) 270 Lake Michi gan 5.6 c lindraceurn - -efi sh) Osmerus mordax who1 e plankton and 3.9t0.4 (16) 0.92tJ0.08 (53) d 4200 Lake Mi chi gan 5.6 (smelt) flesh benthic feeder 0.27tO. 05 (9) 0.92tO. (53) 290 Lake Mi chi gan 5.6 - - 08 d Perca f 1 avescens flesh parti a1 ly 0.3740.03( 24) 0.9250.08 (53) 400 Lake Mi chi gan 5.6 -mTwm pisci verous d flesh pi sci verous 0.2320.03 ( 20) 0.9220 a 08 ( 53) 2 50 Lake Michigan 5.6

Salvelinus flesh pisci verous 0.27tp.02 (24) 0.92+0.08( 53) 2 30 Lake Mi chi gan 5.6 bi"ut, Salmo trutta flesh pisciverous 2 50 Lake Mi chi gan 5.6 (brown trout} 134

L W aE >a, 3 z.

I L nW G c1n ma, c -7

a, W03 E

rln -r- rc El 135

Table 3.5.2 Concentration and coefficients of variation of manganese concentration ( ppm) i n water and freshwater fishes (fresh weight).

Water Flesh Whole Fishes

Arithmetic Mean 0.028 + 0.034 0.34 + -17 7.9 + 5.5 Coefficient of Variation 11% 5a 7m N 7 16 14 Geometric Mean 0.014 x- 4.43 0.32 x 1.45 6.7 1.84 Coefficient of Variation 343% 4 5% 84% 7 14 N 16 136

Using the gsotnetric means of stable manganese concentration in freshwater fishes in Table 3.5.2 and Equation (2.1.3),

5i-'(Mr1)~= 6.?/[Mnlw (3.5.1) for whole fish and

(3.5.2) for fish flesh, where [MnIvJ has units of ppm. These relations and data from Table 3.5.1 are plotted in Figure 3.5.1. Note that it is inconsequential whether filtered or unfiltered values are plotted on the graphs However, for predi cti ve purposes, an unfil tered value for the manganese concentration in water should be used (Section 2.5.3) unless the user is willing to accept the gross overestimates obtained

by USE? of fi1 tered concentrations )I We hypothesize that the differences in manganese concentration among species of fishes are owing ta physiological demands rather than differences in concentration of manganese in the fishes' food. However, bottom feeders have lowest concentrations, plankton feeders have the highest, and piscivores have intermediate concentrations. These differences ineri t further research.

3.5.3.2 Manganese Bioaccumulation Factors for Plants

Table 3.5,3 sunimarizes the few data available ~n manganese bioaccumulation factors for freshwater plants. Harvey (1969, 1970), using a continuous flow, laboratory system to control temperature, showed that nonlethal temperatures have no effect on the manganese bioaccumulation factors for algae. On the basis of data in Table 3.5.2, we recommend 137

CONCENTRATION IN WATER (ppm) Figure 3-5,: Bioazcumulation factors for manganese -in freshwater fishes as a function 05 manganese concentration in water, Table 3.5.3 Bioaccumuiation factors for rianganese in aquatic plants

Stable element Stable element 54Mn 54Mn Species Description coricen trati on concentration concentration concentration BF~ Location Reference in p3ant in water ( m1 in plant in water (?Prnj Filtered Un!P!tered (pCi/gj (pCi/l)

phy top 1an k ton 1154 (?4!b 0.003922 '2,090 Lake Michigan 5.5 0.00008 (53)

Mix of 1. Lynqbya, 1. filamentous 11.0 0.48' 23,000 Reactor discharge 5.1 Oscillasoria, Phormidium; blue-green algae; canal on 2. Cosmarium; and 3. 2. green algae; Connecticut River ______Njtzchia, Melosira and 3. diatoms Stigeocionium 1tibrjcun filamentous green 2.3~105 4.5x1a4 5,000 Laboratory, contindous 5.12 a1 gae flow sxp. (23"-32OC) Navicula seminulum unicellular 0.70~10~ 4.5~10' 1,600 Laboratory, continuous 5.12 diatom flow exp. 2 (23O-32"C) W 5 4 03 Plectonema boryanum filamentous 1.4~10 4.5xj0 3,200 Laboratory, continuous 5.12 bl ue-green a4 gae f?or/ exp. (23O-32Oi) submerged macro- 250-2590 25,00Gd Hudson River 5.:4 Potamogeton perfs1 iatus 0 .01-0.1 phyte Ceratophyllum sp. rootless, sub- ai 0.003 0.01 8,100 Lake Maggiore, 5.19 merged thin Ita1 y , me sotrop h i c stemmed macrophyte Lagarosiphon sp. 6 0.003 0.01 600 Lake Maggiore, 5.10 Italy, nesotroph

Table 3.5.3 (continued)

Stable element Stable element 54Mn 54Mn Species Description concentration concentration concentration concentration BF~ Location Reference in plant in water ( m) in plant in water ( PPm) FiTGZT-dkered (pCi/g) (pCi/l)

mostly submerged, 32 0.026 1,200 Lake Maggiore, 5.17 top of plant out Italy, mesotrophic of water emergent 28 0.026 1,100 Lake Maggiore, 5.17 Italy, mesotrophic Nymphea lutea large floating 9 0.003 0.01 900 Lake Maggiore, 5.19 (water lily) 1eaves Italy, mesotrophic submerged 13 0.026 540 Lake Maggiore, 5.17 t%2d 1 Italy, mesotrophic large floating 5 0.026 190 Lake Maggiore, 5.17 leaves Italy, mesotrophic

--I a8ioaccun~ulationfactors are all based on unfiltered concentrations in water unless only filtered values were reported. w u3 bMean S.E. (no. of samples) of all non-zero "corrected" (Copeland and Ryers 1970) concentrations.

'istimate from release rates. dEstimte fron nonlinear regression analysis. 140

4 that a value of 10 be used as the manganese bioaccumulation factor in algae. For submerged macrophytes we recommend a bioaccumulation 4 factor of 10 . On the basis of the fm$ data from Lake Maggiore, we 3 recommend that a value of 10 be used for macrophytes with floating leaves and emergent vegetation.

3.5,3.3 Manganese Bioaccumulation Factors for Invertebrates

3.5.3.3.1 Mal-sganese Bioaccumulation Factors for Molluscs In bivalves manganese concentrations in both the shell and soft t‘issues increase with size (Merlini , 1966; Merlini el; al e , 1965; Merlini et al . , 1967). This fact explains part of the standard error associated w-i th the mean concentrations of manganese given for bivalve tissues in Table 3.5.4. The data in Table 3.5.4 were averaged over a1 1 size classes. Manganese concentrations in tiie mantle and visceral sac may vary considerably with nature of the substra%e for the same lake. This factor is in large part responsible for the large standard

errors far I__-Un-a’o in Lake Maggiore since IJnio was taken from sites having different substrates - In snails, manganese concentrations do not vary greatly with size

(Merlini, 1966; Merlini et a1 , 1965; Merlini et a1 , 1967). Moreover, only mal 1 differences in manganese concentration were found in the snail _.-.IViviparus .-- with change in habitat within the same lake (Merlini

-. et a1 . , 1965). I he data on bioaccumulation factors fer manganese in niolIiiscs (-Table 3.5.4) suggest that EF(PI~)~is constant or that manganese concentrations are only partially homeostatically contr-olled. the bioaccumulation factor for Unio tissues in Lake Maggiore is about Table 3.5.4 Bioaccumulation factors for manganese in molluscs

~ ~ Stab1 e element Stable element concentration concentration Taxon/Tissue in tissue in wateg Bi oaccumul a tion Factor Location Reference twm! t ppm)

Bivalves : Anodorita cygnea Lake Maggiore 5.18 Shel 1 41 9+50( 2)b 0.016 26,000 Gills 2450F150(2)b 150,000 Mantle 720519@(2Ab 45,000 Visceral sac 155+35(2) 9,700 Muscle 30.5c12( 2) 1,900 Anodonta cygnea Who1 e 341 236 ( 4 ) $0.1 3,400 Irrigation canal, 5.4 Crescentino, Italy Whole 20627 (4) 2.5.1 2,100 Irrigation canal, 5.4 Vprrplli: Jtaly Lamell idens marginal is 0.@1lC Lake Masunde. India 5.21 Labial palps 987 2 900,000 Gills 321 0 300,000 Mantle 1228 110,000 Visceral sac 1220 110,000 Muscle 834 80,000 Foot 3 54 30,000 Total soft 1247 110,000 Shel 1 942 90,000 Qu;I;I?~ pustulosa 0.004' 5.20 48 6 120,000 -____Unio mancus elongatus 0.016 Lake Maygiore 5.16, 5.18, 5.19 Shel 1 3&2+43( 1 2) 2u,ooo Gills 2 1 0@7260( 7 ) 130,000 Mantle 14OG335( 7)d 88,000 Muse1 e 26s53 (7 16,000 Viscera1 sac 37E96 (7Id 23 * 000 Table 3.5.4 (continbed)

Stable element Stable element concentration con cell t rat ion TaxonlTissue in tissue in water Bi oaccumu’lati on Factor Location Reference IPPd (PPmIa --Unio mancus elongatus 0.033 Varese 5.16 She1 1 57 e57( 3 j 26,000 Mantle 2400+300 (3’ 7; ,oco Visceral sac 6(3@58( 318 18,000 Snails: 0.016 Lake Magyiore 5.18, 5.19 35+1(4) 2,200 za~i(4) 1,500 Lymnea peregra Whoje 226241 (3) ’Lc. 1 1,303 Irrigation canal, 5.4 Casalino, Italy Whol e 85 (1) 4.1 850 Irrigation canal, 5.4 Vercelli, Italy Lynmnyestagna1 is 1 0755 ( 6 ) 4.1 1,iOO irrigetion canal, 5.4 Casalino, Italy Uhol e 7 026 ( 5 ) ’L2.1 703 Irrigatl’on canal, 5.4 Crescent ino, Italy Whole 1021 4( 15) so.1 7,103 Irrigation canal, 5.4 Vercelli, Italy Ph sa acuta -klT 2552 6 1 0.009 Irrigation canal, 5.4 (I 2,800 Cameri , I tCly Whol e 85513(2) %O.l 850 Irrigation canal, 5.4 Cjsalino, Italy Planorbis sp. Whol e 599+221(3) ‘Lo.l 6,900 Irrigation canal, 5.4 Casalino, Italy Whol e 750+_56(2) ‘Lo.l 7,500 Irrigation canal, Vercelli, Italy 5.4 I43

d; -?=? In mln

"3.,

WL 0 .r 0E .r m0, +J m V 2 -10 Yal m ^1

L- +JO V LLm S 0 .C c, 00 0 %a OLn 0 r-- Lor- 0. Ls in N u5 mu 0 .r M

ID 9- P 0 0 2

-0 T5.G mr- 7- .--.w -W v m 14 Ln m0+1+1 +I cor-coN N N0

3W VI VI .r t- --.S 0X

I- 144 the same as that in Lake Varese despite a two-fold difference in manganese concentration in the water. The whole-body bioaccumulation factor for Physa is about 3.4 times higher for the environment having a inanqanese concentration in water an order of wagnitude lower than that of the other. However, we must not put too much emphasis on these resul ts because the manganese concentrations in water, which are based on only a few measurerwnts, can only be considered rough indicators of the concentration that the organism would be exposed to over long periods of time since the manganese concentration in water has strong seasonal variability (Section 3.5.1). Further, as pointed out in

Section 2.5.5, sediment contamination may affect whole-body concentrations af manganese in invertebrates.

For studies in which bioaccurnulation factors are based on unfiltered values for the manganese concentration in water, the bioaccumulation factors for all soft tissues in bivalves except gills are less 5 5 than 10 . Thus, \re recommend that a bioaccumulation factor of 10 be used for soft tissues of bivalves (Table 3-5.4). Note that the value of this recommended bioaccurnulation factor compares favorably with the 4 54Mn bioaccumulation factor of 4.5 x 16 determined from the fallout data of Gaglione and Ravera (1964) for all soft tissues of Unio from

Lake Maggiore. On the basis of data in Table 3.5.4, we recommend that 4 a bioaccumulation factor of 3 x 10 be used for bivalve shells.

Based on the few data for snails we recommend that a bioaccumulation factor of 184 be used for snail shells and whole snails, For 3 soft tissues of snails, we recommend a bioaccumulation factor of 2 x 10 . 145

3.5.3.3.2 Manganese Bioaccumulation Factors for Invertebrates other than Mol 1 uscs

On the basis of data in Table 3.5.5, we recommend that a biaaccumu- It apparent from lation factor of 70 4 be used for crustaceans. is Table 3.5.5 that there is considerable variation in manganese bioaccumulation factors among insect species. Owing to the lack of comprehensive 3 data, we must recommend that a bioaccumulation factor of 10 be applied to insects in general. Table 3.5.5 Eioaccumulatiori factors for rwiqanese in invertebrates other than molluscs

Stable element Stable element Size class or concentration concentration Species Descr.i p t ion l-ife stage iissde in tissue in water BF Location Reference (pprC) (P?b)a

Calanoid arid cyclopoid zoopi an kton whol e .9250.08( )' Lake Mich'igan copepods whol e C 53 3,900 5.5 4,600 Lake Mich-igan 5.5 Amphi poda (mostly small shrimo whol e 7.452.1 ('I7) 0.92TO. ( Lake Michigan Pontoporeia sp.) 08 53) 8,000 5.5 Ranatra linearis carnivorous acuatic adu': t whole e 51.8 (1) 9 6,000 Caneri, Italy 5.4 heniipteran, feeding adu: t wh3: e 333.m(2) %lo0 3,030 Casaiino, Italy 5.4 on tadpoles, insects a31;lt wnol e 3545s(2 ) Ql oc 4,OGO Crescenti no, I tj'iy and liirvae 5.4 adult whole Notonecta glauca aauatic insect 6.910 .S( 2) %loo 70 Casal ino, I ta'ly 5.4 whol e 5.8- ('I) ""'I 00 60 Oldenico, Italy 5.4 Hydroph.ilus piceus carnivorous aquatic adult whole (2) 9 1,000 Cameri, Italy insect e 9.420.7 ""100 Casajico, Italy 5.4 who1 34.457.6(4) 5.4 A rrnol e ""1 300 Crescentino, Italy -P 00 830 5.4 cn whol e it:g:; [; 1 2.1 00 Olber;ico, Italy 5 .c rrnol e 2 2.1 400 Vercel: i Italy .j 7.3TG. (4) oc; 200 , 5.4 Dy s ticils ma rg i ria 1 i s ac;l;atic insect adult whol e 6.6 (1) 9 7CO Caineri, Italy 5.4 whole 14.7i7.9(2) %-I 00 100 Casalino, Italy 5.4 whol e 40.1 (1) %loo %00 Cresien-tino, Italy 5.4 parasitic leech whole ?.loo 100 0'1 deni co, I ta'ly found on fish 9.5 5.4 and insects aUnfiltered values Liniess indicated otheyise. bMean 2 S.E. (no. of samples) of all non-zero "correcLed" (Copeland and Ayers 1970) concentrations. 'Filtered value. %ncorrected va: ue. e- irrigst>or: canal at this and followins locations. 147

References for Section 3.5

The numbers in the left-hand margin correspond to the reference numbers used in the tables of section 3.5.

5.1 Blanchard, R. L., and B. Kahn, The fate of radionuclides released from a PWR Nuclear Power Station into a river, IN Symposium on environmental behavior of radionuclides released in the nuclear industry, Aix-en-Provence , France, 14-1 8 May , 1973, IAEA/SM-172/26.

5.2 Bortoli de, M. C., P. Gaglione, 0. Ravera, and G. de Fisica, Fall- out manganese-54 and zinc-95 in water and fishes of four lakes in Northern Italy, Minerva Fisionucleare -13, 72-77 (1969). 5.3 Bryan, G. W., and E. Ward, The absorption and loss of radioactive and non-radioactive manganese by the' lobster, Homarus -vulgaris, J. Mar. Biol. Ass. U.K. 45, 65-95 (1965).

5.4 Cavalloro, R., and M. Merlini, Stable manganese and fallout radiomanganese in animals from irrigated ecosystems of the Po Valley, ECOl Ogy -48, 924-929 (1 967). 5.5 Copeland, R. A., and J. C. Ajers, Trace element distribution in water, sediment, phytoplanktan, zooplankton, and benthos of Lake Michigan: A baseline study with calculations af concentration factors and buildup of radioisotopes in the food web, Environmental Research Group Special Report Number 1, Great Lakes Research Division, Univ. Michigan, Ann Arbor, 1970.

5.6 Copeland, R. A., R. H. Beethe, and W. W. Prater, Trace element distribution in Lake Michigan fish: A baseline study with calculations of concentration factors and equilibrium radioisotope distributions, Environmental Research Group Special Report No. 2, Great Lakes Research Division, Univ. Michigan, Ann Arbor, 1973.

5.7 Cross, F. A., and J. H. Brooks, Concentrations of manganese, iron, and zinc in juveniles of five estuarine-dependent fishes, IN Radionuclides in Ecosystems, 0. J. Nelson (ed.), pp. 769-775, Proc. of Third National Symposiuin on Radioecology, Oak Ridge, Tennessee, Play 10-1 2, 1971 { 1973). 5.8 Cross, F. A., L. H. Hardy, N. Y. Jones, and R. T. Barber, Relation between total body weight and concentrations of manganese, iron, copper, zinc, and mercury in white muscle of bluefish (Pomatomus sol tatrix) and a bathydetnersal fish Antimora rostrata, J. Fish7 Res. Bd. Can. -30, 1287-1290 (1973). 14%

References for Section 3.5 (continued)

5.9 Gag1 ione a P. , and 0. Ravera Manganese-54 concentrations in fa1 1 - out, water and Unio mussels of Lake Maggiore, 1960-53, Nature -204, 1215-1216 (1964).

%,lo Gibbs, R. J., Mechanisms of trace metal transport in rivers, - Science 180, 71-33 (1973). c.11 )Harvey, R. S., Effects of temperature on the sorption of radio- i, ’ nuclides by a blue-green algae, IN Symposium on Radioecology, J. Nelson and F. 6:. Evans (eds. 1 , pp. 266-269, Proceedings of theD. Second National S posium on Radioecology, Ann Arbor, Michigan, May 15-17, 1967 (196931: 5.12 Harvey, R. S., Temperature effects an the sorption of radionuclides by freshwater algae, Health Phys. 19, 293-297 (1970).

5.13 Wutehinson, E., A treatise on limnology, Vol. I, John Wiley and Sons NewG. York , 1957.

5.14 Lentsch, 3. W., T. J. Kneip, M. E. Wrenn, P. Howells, and M. Eisenbud, Stable rrianganese and manganese-54G. distributions in the physical and biological components of the Hudson River Estuary, IN Radionuclides in Ecosystems, D. J. Nelson (ed.), pp. 752-768, Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tennesseeg May 10-12, 7971 (1973).

5*15 Marshall, J, S., and J. ti. LeRoy, Iron, manganese, cobalt, and zinc cycles in a South Carolina Reservoir, IN Radionuclides in Ecosystems, D. J, Nelson (ed.), pp. 445-473, Proceedings of the Third National Syiposiurn on Radioecology, Oak Ridge, Tennessee, May 10-12, 1971 (1973).

5.16 Merlini, M., The freshwater clam as a biological indicator of radiomanganese, IN Radioecological Concentration Process, B. Abery, and F. P. Hungate (eds.) , pp. 977-982, Proceedings of the International Symposium, Stockholm, April 25-29, 1966.

!ie17 Merlini, M., C. Bigliocca, A. Berg, and G. Pozzi, Trends in the concentration of heavy metals in organisms of a mesotrophic lake as determined by activation analysis, IN Nuclear techniques in environmental pollution, pp. 447-458, MEA Symposium on the Use of Nuclear Techniques in the Measurement and Control of Envir- onmental Pollution, Salsburg, Austria, Oct. 25-30, 1970 (1971).

5.1% Merlini, M., F. Girardi, R. Pietra, and A. Brozgelli, The stable manganese content of molluscs from Lake Maggiore determined by activation analysis, Limnol . Oceanagr. -_10, 371 -378 (1965). i49

References for Section 3.5 (continued)

5.19 Merlini, M., F. Giradi, and E. Pozzi, Activation analysis in studies of an aquatic ecosystem, IN Nuclear Activation Techniques in the Life Sciences, pp. 615-629, Proceedings of the Symposium held by the IAEA in Arnsterdarri, May 8-12, 1967, also STf/PiJB/155, SM-91/7 (1967).

5.20 Nelson, D. J., T. C. Rains, and J. A. Norris, High purity calcium carbonate is freshwater calm she1 1 Science .I- 152, 1368-11 370 (I 966). 5.21 Patel, E., and G. R. Doshi, Lamellidens marginalis L. as indicator

of manganese-54, Archivio di-Oceanografia e Limnol ogic I17, 27- 42 (1972). 5.22 Schroeder, H. A., J. J. Balassa, and I. H. Tipton. Essential trace metals in man: Manganese, J. Chron. Dis. -16, 545-571 (1966).

5.23 Wangersky, P. J., Manganese in ecology, IN Radioecology, V. Schultz and A. W. Klement, Jr. (eds.), pp. 499-508, First National Symposium on Radioecology, Sept. 10-15, 1961 (7963).

3.6 COBALT

3.6.1 Cobalt Metabolism

Cobalt is essential for some bacteria, some fungi, several species of blue green algae and several species of mammals (Bowen, 1966). Cobalt I1 activates enzymes. A principle use of cobalt I11 is in cyanocobalamin (vitamin B12). Cobalamin is not synthesized by animals but is synthesized by bacteria and actinomycetes (Sherman, 1957). Cobalamin is required by some green (algae and diatoms, many dino- flagellates and yellow-green algae, iigher plants, insects, fish, birds, and mammals.

Absorption efficiency of cobalt from food is about 50% in snails (Nelson and Malone, 1969) and about 5% in fishes. These low absorption efficiencies are in part responsible for the decrease in the cobalt bioaccumulation factor with increasing trophic level that is sometimes observed (Kevern and Griffith, 1966; Lowman, 1963; Morton, 1965; Ophel and Fraser, 1973). This is because the ratio of intake rate to elimination rate gwerns isotopic concentration of cobalt in the organism [Equations (2.4.3) and (2.4.4)].

Elimination rates for cobalt are given in Table 3.6.1. Good estimates of turnover rates come froin studies in which the organisms have come into steady-state with the source of radionuclide. Jn such studies, long time components contain most of the radionuclide. Because of its relatively short radioactive half-1 ife (72 days),

58C0 will have a lower bioaccumulatian factor than the bioaccumulation factors for 6oCo (5.27 year half-life) and stable cobalt. However, as can be seen from the biological half-times given in Table 3.6.1, 0 ab le 3.6.7 ~.~o’iogicaIelimination rates ?or “Cs in var.a’aus aquatic organ-isms

Lemna minor macrophyte 19 from elimination 100 6.27 Tckweedj 3.6 after five month up .ta k e Vallisneria americana macrophyte 51 1.4 from e 1 i mi nati on 100 6.27 (wi I dclier after five nonth uptake 51 1.4 FI-GF, elinination 90 6.30 after five month uptake snail 4.5 15. fast compartment 25 6.28 -isi 1 e formi L av a s 320 0.22 slow compartnent 25 6.25 snails tagged by d cn short term feeding N Larn~s 1 radiata am i is c7 4.8 14.4 f as t cotrip a r tnen t 51 6.14 277 0.25 slow cornpartrnent 49 6.14 Cambarus lonaulus c ray f i h s 70 0.99 5 g atiimls, high 90 6.35 ’level dcse 37 1.9 6.9 cj animals, high 90 6.35 level dose er,i cty4 us nevrt 158 0.44 f.i ve month uptake 6.30 Gi vi ri des cens 90 Ictal urus me1 as fish 1.5 46. single feeding 93.9 6.32 (b’lack bu! 40.5 1.7 single feeding Ihead] i.8 6.32 long, undetsrmined single feeding 6.32 undetermined 0.3 3.5 20. 4 days of water uptake 50 6.32 3.5 2 days of water uptake 35 6.32 .Q 4 6.32 4 ong 9 undete rmi ncd 4 days of water uptake 15 undetermined sp. Insect 135.3 0.66 uptak.2 from water tas 89.3 5.13 Tri choptera 38.4 1.6 uptake from food 90 6.13 larva ingestion 153 the 58Co bioaccumulation factors are not expected to be greatly lower than the bioaccumulation factors for 6oCo or stable cobalt in many organisms. Considering the uncertainty in both biological half- times and in the bioaccumulation factors for- 60Co and stable cobalt, we will not distinguish between the bioaccumulation factors of

58C~and the 1onger-l ived isotopes.

3.6.2 Environmental Cobal t and Ava-*labili ty

The proportion of cobalt in the particulate phase varies greatly among bodies of water. In Par Pond, an oli~otroph~creservoir of law turbidity, 22% of the stable cobalt in the water is in the particulate phase (Marshall and LeRoy, 1974). Greater than 98% of the stable cobalt in the Amazon and Yukon Rivers is particulate. The cobalt appears as adsorbed on organ;c solids and in the crystal structure of sediments. Fukai and Murray (1973) contrast the flractisn of radiocobalt appearing in the particulate phase in the Columbia River to that in the Clinch River, Tennessee; the Columbia had from 80 to

95% (at Vancouver, Washington) and .:he eutrophic Clinch, from 2 to

302, despite the much higher suspended solids concentration in the latter- river. In Perch Lake, a small d~strophic-ew$Ir,aphiclake, essentially all the radiocobalt in the water is in the dissolved phase (Ophel and Fraser, 1973) in contrast to mesotrophic Lake

Maggjore in which roughly 80% of stable cobalt is in the particulate phase (Merlini et a]., 1967). These observations would suggest that the proportion of cobalt in the particulate phase increases with increase in suspended solids concentration and decreases with 154

increase in eutrophy. This latter correlation may be related to the tendency of cobalt to form associations with dissolved organic matter. Thus, increased dissolved organic matter concentrations in eutrophic waters may serve to keep radiocobalt in solution. As indicated in Section 2.2, the soluble forni of a radionuclide is generally more available to algae and subsequent trophic levels ,than the particulate form of the radionuclide. Thus, it would seem that the greater the proportion of soluble cobalt, the greater the availability. However, as indicated in Section 2.2, chelated radionuclides are less available to food webs than radionuclides appearing as free ions. For this reason, soluble cobalt in eutrophic environments is probably less available than soluble cobalt in oligotrophic or mesotrophic environments due to the tendency of cobalt to chelate or form other associations with dissolved organic matter. This expected pattern is seen in the review of the literature on cobalt bisaccumulation factors which fol laws.

3.6.3 Review of Cobalt Bioaccurnulation Factors

Because most investigators have based their bioaccumulation factors on filtered concentrations of cobalt isotopes in water, we generally report filtered bioaecumulation factors. As indicated in Section 2.5.3 use of filtered bisaccumulation factors should generally lead to overprediction of radionuclide concentration in organisms. However, we noted that in nonturbid waters and in eutrophic waters, most of the radiocobalt in the water would be 1 55' in the soluble phase. In these wat?rs, use of filtered bioaccumulation factors would not overpredict radiocobalt concentrations in organi sms .

3.6.3.1 Coba t Bioaccumulation Factors for Fishes

Res ea rc h rs have reported cobalt bioaccumulation factors for flesh and whole bodies of fishes. The relative concentrations of 6oCo in black bullheads from Irlhite Oak Lake may be used to convert these bioaccumulation factors to biDaccumulation factors for other tissues. The relative steady-state concentrations are given in Table 3.6.2. The organ with the highest cobalt concentration is the kidney, which has a concentration 23 times that of the whole body. Note that the flesh to whole body ratio is 0.3, which compares well with the ratio of 0.37 determined for brown bullheads from Perch Lake, Ontario (Ophel and Fraser, 1973). All bioaccurnulation factors for cobalt in fishes were derived from steady-state concentrations of 6oCo or stable cobalt in natural bodies of water. Because cobalt bioaccumulation factors are expected to decrease with increasing eutrophy, the trophic states of the bodies of water were recorded. The bodies of water fell into two categories, eutrophic and mesotrophic; bioaccumulation factors for the former are given in Table 3.6.3, and bioaccumulation factors for the latter, fn Table 3.6.4. Mean bioaccumulation factors for whole fishes and fish flesh are given in Table 3.6.5. As expected, lower cobalt bioaccumulation factors are seen for eutrophic environments. We recommend use of the bioaccumulation factors given 156

Table 3.6.2 Relative steady-state concentrations of “CQ in tissues of ______Ictalurus _____melas (black bullheads) from White Oak Lake, Calculated from data of Reed (1971)

__.---_1_ .- Tissiie Concentration

l...ll -.l_l.-____l_..- ’rissue ^-I ...... Whole Body Concentration ..... 131 ood 4.3

Skin 0.92

F’o esh 0.38 5 2 Liver 3.1

Gut 4.0 Kidney 23 '. I.

Table 3.6.3 Bioaccumulation factors for cobalt in fishes from eutrophic environments (based on filtered water concentrations)

Stable element Stable element Tissue concentration concentration Location Reference Species Size(in.) Description in tissue in water BF IPW) ( PPb 1 Perca flavescens Achl t pi sei vorous who1 e 18 Perch Lakc 6.29 (yellon perchr (8.7-13) Canada t Adcil flesh 9 6.29 young zooplankton and whole 130 6.29 (2.4-3.9) insect 1arvae I cta 1Llrus Advl t omnivorous , whole 52 6.29 nebul os us (6.7-11.8) pi ant material (brow bul Ihead) Adult flesh 19 6.29 Young whole 63 6.29 (2.0-2.8) Ariii? t vorow omni , plqt, whgje 80 6.29 (C*7-7.9) young eating insect whole 94 6.29 (2.0-3.2) 1 arvae ___-Seil-oti !us Adult unknown whole 18 6.29 var5at-i ta (2.8-4.7) Ipcarl dace) Adult unknown whole 20 6.29 (2 .O-3.9)

Adul t unknwn whole 38 5.29 (2.0-3.9) ..- Dorocora piscivorous whoie fish ! 32 White Oak 6.26 ciipecianurn less Lake, :gizzard shad) viscera Tennessee 6.26 Lepocis mcro- chi ronomi d whole fish , 26 -3z-FX Ibiuegill) larvae, ter- less restrial insects, viscera fish, carnivorous 158

aJU C h L aJ P Q-aJ UJ czaJ

W w C 0 .P+J rtl U 0 -I

In IA- f a v)

ar P-o .l= 3

cn VI -.tu- VI VC cl QP CU- a-m= N LW -7 ma vl c

VI 01 .r V V al 4 a v) .@V Tzble 3.6.3 (continued)

Stable element Stable element Size class Species Tissue concentration concentration BF Location Reference (in.) Descriptfon in tissue in water

Assorted fishb sight feeders flesh L-.r'5J 6.5, 6.25 --Esox lucius pi sci vorous flesh (0.03-0.11)c 3 -6f (northern pike) 0.07 (1 23 IllinsisRiver 6.22 b!i cropterus piscivorous flesh (0.06-0.18) 30 6.22 s al~oides 0.09 (=th bass ) piscivorous flesh 0.07 (0.05-0. 0) 23 6.22 Le 8) isostdus piscivorous flesh 0.15 (0.11-0. 50 6.22 +S +S !+bar) bii cropterus f? 6.22 -1 omi esh e.?s(o.?4-0.16! 50 Lii do(mal 1 lmouth eui bass) Ictiobus omnivorous flesh 0.081 (0.06-0.1 2) 27 6.22 cwri ne1 insect larvae, i g thlus ( b n:ou 1us ks , a1 gae, buff a10) molaquati.c plants Dorosoma omnivorous 6.22 -- flesh 0.16(0.1-0.25) 53 i risect' 1arvae, =had) mol?usks, algae, aquatic plants

Moxostoma macro- omni vorous flesh 0.083(0.05-0.11) 28 6.22 insect larvae, %%%--- mollusks, algae, redhorse) aquatic plants N 9 u3

a 1, ,

Table 3.6.4 Bioaccumulation factors for cobalt in fishes from mesotrophic environments (based on filtered water concentrations)

Stable element Stable element Species concent rat1on concent rati on Location Reference Size class Description Tissue in tissue in water BF [in.) lpwl (ppb) (filtered)

losa Eeudoharen mop1 ankton feeder whole 6.029 -L 0.014 0.1920.02 190 Lake Michigan 6.6, .6;20 +ew i us 0.6-0.8 A1 os 3 Eeudoharengus 4.5-70 zooplankton feeder whole 0.065 +- 0.005 0.19 2 0.02 420 Lake Michigan 6.6, 6.7 -mew 1 f e 1 benthic feeder Osrerus rmrdsx zooplankton feeder whole 0.19 ? 0.05 0.19 k0.02 ,ffi?O\ (sr;lelt) 5 Yl-1 A1 burn us alborel 1 4.3-5.1 plankton feeder flesh 0.072 0.002 0.02 600 Lake Maggiore 6.23,6.24 ---(-ma) a I

Coreqonbs 9.8 plankton and benthic flesh 0.0077 0.0016 0.02 Lake Maggiore 6.23,6.24 A feeder f 385 2 -”del la) Osmerus mordax plankton and benthic flesh 0.043 ? 0.004 0.19 t 0.02 230 Lake Michigan 6.6, 6.7 (sc,el t]- >5 feeder 6.6, whole 0.091 f 0.011 0.19 f 0.02 480 Lake Michigan 6.7 or;isccnaycus 2.4-3.7 benthic whole 0.024 0.002 0.19 r 0.02 130 Lake Michigan 6.6, 6.20 *perch)Perccqsis f Lake Michigan 6.6, 6.20 Notrcc hudsoni ~s 2.2-2.5 benthic and zoo- whole 0.042 +- 0.015 0.19 2 0.0 220 !sp 0 tta-fr-i iinzr) plankton whole 0.12 0.03 0.02 Michigan 6.6, 6.7 Notropis hudsonilrs 4.5-5.5 benthic +- 0.19 2 630 Lake ispottail snitier) flesh 0.041 0.003 0.19 0.02 Mtchigan 6.6, 6.7 Notrc is hudsonius 7.5-8 unknown t k 220 Lake T-C-spotta? 1 shiner) .-_Sczrdinius erythro- 1 omnivorous flesh 0.0058 kO.OOl1 0.02 290 Lake Maggiore 6.23,6.24 p b til a 1!iius(ruda’r flevescens 8-13 partial ly flesh 0.053 f 0.0089 0.19 20.02 280 Lake Michigan 6.6, 6.7 fYE17LiFiGSgPerca piscivorous h

WW 2s c *I- I-' c 0 U v d

ID m W P- a' t-a 163

Table 3.6.5 Mean bioaccumulation factors for cobalt in fishes from mesotrophic and eutrophic waters

Bioaccumul ation factors Nature of Number of body of water bodies of water F1 esh Mho1 e

Mesotrophic 3 323 L 47(11) 439 k 115(7)

Eutrophic 4 26.6 4 3.5(16) 43.8 k 10.4(14) 164 in Table 3.6.5 for mesotrophic and eutrophic environments. In the absence of data for oligotrophic waters, we suggest using the mean value from the mesotrophic waters. Lucas and Edgington (1970) have suggested that cobalt concentration may be under partial homeostatic control. The inverse correlation between the cobalt bioaccurnulation factor and eutrophy could be due to partial homeostatic control if cobalt concentration in the water increased with eutrophy. In Figure

3.6.1 the cobalt concentrations in fish flesh are plotted against cobalt Concentrations in water for three mesotrophic waters and one eutrophic water, the Illinois River. If the slope (exponent) of the logarithmic regression were I, the cobalt bioaccumulation factor would he independent of the cobalt concentration in water.

Sitice the slope is less than 1, either partial homeostatic control or eutrophy coincident with increasing cobalt concentration in water may be operating to diminish the slope. Lack of.more exten- sive data on bioaccumulation factors from environments exhibiting a wide range of cobalt concentrations within each of a few distinct eutrophy categories prevent us from firmly deciding between the hypotheses of partial homeostatic control and eutrophy. In view of the known tendency of cobalt to form organic associations, the latter explanation seems more plausible. Some workers have observed that the cobalt bioaccumulation factor in fishes decreases with increase in tropkic level (Morton, 1965; Nelson et a1 . , 1971; Ophel and Fraser, 1973). Other workers have not found a clear-cut correlation (Kevern and Griff i th, 1966 ; I.

ORNL-DWG 74-7032

0.01 0.02 0.05 0.1 0.2 0.5 1 2 5 STABLE COBALT CONCENTRATION IN WATER (ppb) Figure 3.6.1 Concentration of stable cobalt in fish flesh in various environments. 166

Mathis and Cummings, 1973). For this reason we have specified the

feeding habits of the fishes in Tables 3.6.3 and 3.6.4. Based on the limited and conflicting data at hand, however, we cannot specify a relation between the cobalt bioaccumulation factor in fishes and their feeding habits.

3.6.3.2 Cobalt Bioaccumulation Factors for Plants

In Table 3.6.6 are listed cobalt bioaccumulation factors for algae and vascular plants. Compared to fishes the cobalt bioaccumulation factors of plants are high. On the basis of the few data on algae in Table 3.6.6, we reconmiend a cobalt bioaccumulation factor of 104 for algae. The two species of emergent vascular plants exhibit widely differing cobalt bioaccumulation factors. The low bioaccumulation factor was calculated from ‘OCo concentrations , and the high bioaccumulation factor was calculated from stable ele- men% concentrations. Possibly, because diffusion of elements in interstitial water of sediment is slow, the ‘OCo concentration

in interstial water from which ...Pontederia obtained its 6oCo had not reached steady state with the ‘OCo in the water column above. Based on the stable cobalt bioaccumulation factors for Phragmi tes , we recommend a bioaccumulation factor of 103 for emergent vascular plants. Physical form of aquatic vascular plants may affect the cobalt bioaccumulation factor. Merlini et al. (1971) reported that floating leaf species have lower bioaccumulation factors than submerged species.

Ophel and Fraser (1973) noted that the finely divided submerged species had the highest bioaccumulation factors, that other submerged Table 3.6.6 Bioaccumulation factors for cobalt in aquatic plants

BF BF Taxon Location Reference { f i1 tered) (unfiltered)

A1 gae: bl ue-green a1 gae 30,000 Connecticut River 6.2 - diatoms 1,500 Lake Michigan 6.6 Navicula seminulum (diatom) 1,100-2,000 Laboratory 6.16 Nitella sp. (green alga) 500 Perch lake 6.29 P1 ectonema boryanum 250-620 Laboratory 6.16 ( f i 1amen tous bl ue-green) Sta’ eoclonium lubricum 2,800 Laboratory 6.16 *Srez;nT-

Emergent vascul ar pl ants :

Phragmites sp. 2,000 4 00 Lake Maggiore 6.23 ’ Pontederia cordata 20 Perch Lake 6 -29

Vascular plants floating 1 eaves : Nuphar sp. 800 160 Lake Maggiore 6.23 Nuphar variegatum 200 Perch Lake 6.29 Nyrnphea 1 utea 900 200 Lake Maggi ore 6.24 Nymphea odorata 200 Perch Lake 6.29 P~~am~geto~ 600 Perch Lake 6.29 ~~~am~~etonna tans 500 Perch Lake 6.29 Sparganiurn fluctuans 300 Perch Lake 6.29 168

emwmcn m T\JN NNCV N WUJ.. ... COracD u3

QJ me, Y mw rb mRl --I ZE maJ I:0 YY R3Rl L a, 1-J a

000 000 Be-a n mmnn r-- Lnm NNNN P c-

SL 169 species had intermediate bioaccumulation factors, and that floating plants had the lowest. In Table 3.6.7 we give mean cobalt bioaccumulation factors for submerged and floating vascular plants from Lake Maggiore and Perch Lake. In Lake Maggiore the cobalt bioaccumulation factors for submerged plants are significantly greater t those for floating plants.

In Table 3.6:7 the mean cobalt bioaccumulation factor for submerged vascular plants from mesotrophic Lake Maggiore is significantly greater than the mean bioaccurnulation factor for dystrophic-eutrophic Perch Lake. This conforms to the expected resul t of decreased cobal t bioaccumulation factors with increased eutrophy, We will assume the mean values given in Table 3.6.7 for ..... Lake Maggiore and Perch Lake to be representative of mesotrophic and eutrophic waters, respectively. Thus, the cobalt ~i~ac~u~u~atjon factors for submerged and floating vascular plants in eutrophic 3 waters are 2 x 10 and 4 x respectively. For mesotrophic and

01 igotrophic waters we recommend cobalt ~io~cc~mulati~~factors of 4 3 10 for submerged vascular plants and 10 for floating vascular plants.

3.6.3.3 Cobalt Bioaccumulation Factors for Invertebrates

The few cobalt bioaccumulation factors available for invertebrates are listed in Table 3-6.8, The cobalt b~~a~c~i~u~a~i~~ factor for the bivalve Unio from mesotrophic Lake Maggiore .is dbout 25 times larger than the bioaccumulation factors for other bivalves from the eutrophic Illinois River and the ~y~trophi~-~ti~~~~hi~ Perch Lake. We attribute this difference: to differences in degree of eutrophy. For bivalves in eutrophic environments we recommend 170

Table 3.6.7 Mean cobalt bioaccumulation factors for submerged and floating-leaf vascular plants from Perch Lake and Lake Maggiore (based on filtered water concentrations)

-I Bioaccuinulation factors Significance of Location ...I__ il_l difference between Submerged F1 oati ng plant types

_I^_

Lake Maggiore I 8,900 t 2,300(5) 850 ? 50(2) P < 0.05 Perch Lake 1,600 f 500(4) 360 * 81 (5) P 2 0.10 Signi f icance of differences between 1 akes P < 0.05 P 2 0.06

11.--.- , I,

Table 3.6.8 Bioaccumulation factors for cobalt in invertebrates (based on filtered water concentrations)

Stable element Stable element Taxon concentration concentration Location Reference Tissue Feeding habits in tissue in water BF (PPd (PPb) Bivalves: Amblema pl icata soft filter feeder 0.7 3 2 00 I1 1 inois River 6.2 Elliptio sp. soft filter feeder 330 Perch Lake 6.9 Fusconaia flava soft f i1 ter feeder 1.2 3 400 I1 1 inois River 6.22 Quadrul a quadrul a soft f i 1ter feeder 0.8 3 300 Illinois River 6.22 --Unio mancus soft f i 1ter feeder 0.1 8-0.20 0.02 9,000-10,000 Lake Maggiore 6.24 Crustaceans: Cambarus sp. whol e 1,600 Perch Lake 6.29 msh) d Copepods whol e filter feeders 0.134 0.19 700 Lake Michigan 6.6 Y Nyysis relicta whol e detritus 0.034 0.19 200 Lake Michigan 6,6, 6.10 Insect 1arvae : Argyracti s sp. whol e 23 ,OOOa Columbia River 6.8 Glossma sp. whol e 11 ,OOoa Columbia River 6.8 Hydrobaeni nae whol e periphyton 5 ,OOOa Columbia River 6.8 Hydropsyche cockerelli whol e phytoplankton 15 ,OOOa Columbia River 6.8 Tend i ped i nae whol e detritus 3 ,OOoa Columbia River 6.8 Snails: Amnicola sp. whol e 4,400 Perch Lake 6.29 Sta nico'la soft 9,000a Col urnbia River 6.9 -%hl-=Es 172

Nd W

W L 0 "rcn (7, CD E

Y(u Ij -1

0 0 0, '.B N I 0 0 ? N r

N 0 0

-- Lo 0 I e N CJ

VI 3

.r4J L u 4) TI

+6-, 0 Lo 173

2 a cobalt bioaccumulation factor of 4 x 10 . For bivalves in mesotrophic or oligotrophic waters we recommend a cobalt bioaccumulation factor of 104 . The few data available for insect larvae and snails 4 suggest a cobalt bioaccumulation factor of 10 for both of these groups. The cobalt bioaccumulation factor for tubificids in the eutrophic Illinois River is 500, a value that we will assume is characteristic of eutrophic environments. The few data for crus- 3 taceans suggest a cobalt bioaccurnulation factor of 10 . 174

References for Section 3.6

The numbers in the left-hand margin correspond to the reference numbers used in tne tables of section 3.6,

6.1 Anderson, E, D. and L. I_. Smith, Jr., A synoptic study of 30 fish species from western Lake Superior, Un-iv. Michigan, Agricultural Experiment Station, Technical 5ul letin 279 (1971).

6.2 Blanchard, R. L. and B. Kahn, The fate of radionuclides discharged from a PWR nuclear power- station into a river, Symposium on Environmental Behavior of Radionuclides Released in the Nuclear Industry, IAEA/SM-172/26 (1973)- J. M. ‘Trace elements in biochemistry, Academic 6.3 Bawen, 14. , Press, New York, 1966. 6.4 Carlander, K. D. , Handbook of freshwater fishery biology, Val. 1, Iowa State Univ. Press, Ames, Iowa (1969). 6.5 Churchill, M. A., J. S. Cragwall, Jr., R. W. Andrew, and S. L. Jones, Concentrations, total stream loads, and iiiass transport of radionuclides in the Clinch and Tennessee

Rivers, ORNL-3721, S~rppl 1 (1965) I i 6.6 Copeland, R. A., and J. C. Ayers, Trace elements distribu- \ tions in water sediment, phytoplankton, zooplankton and benthos of Lake Michigan, Environmental Research Group Special Report No. 1, Great Lakes Research Division, Univ. Michigan, Ann Arbor (1972).

6.7 Copeland, R. A., Ti. B. Beethc, and W. Prater, Trace element distributions in Lake Michigan fish, Environmental Research Group Special Report No. 2, Great Lakes Research Division, Univ. Michigan, Ann Arbor (1973).

6.8 Davis, J. J., Accumulation of radionuclides by aquatic insects, IN Biological problems in water pollution, p. 211, U.S. Dept. Heal th, Education and We1 fare , Madison , Wisconson (1955).

6.9 Davis, J. J., R. Perkins, R. F. Palmer, W. C. Hanson, and J. F. Cline, W.Radioactive materials in aquatic and terrestrial organisms exposed to reactor effluent water, Prac. 2nd Intern. Conf. on The Peaceful Uses of Atomic Energy, United Nations, Geneva, Sept. 1-12, 1958, 18, 423-428 (1958). 175

References for Section 3.6 (continued)

6.10 Edgington, D. N., E. Harvey, and M. S. Torreyo A preliminary study of the chemical composition of the opossum shrimp, Mysis relicta, IN Radiological and Environmental Research Division Ann. Report, R. E. Rowland, ANL 7960, Part I11 (1972). 6.11 Friend, A. G., The aqueous behavior of strontium-85, cesium- 137, zinc-65, and cobalt-60 as determined by laboratory type studies in transport of radionuclides in freshwater systems, Conf. in Austin, Texas, Jan. 30-Feb. 1, 1973, TID-7664.

6.12 Fukai, R., and C.’ M. Murray, Environmental behavior of radiocobalt on radiosilver released from nuclear power stations into aquatic systems, IN Syinposium on Environmental Behavior of Radionuclide Released in the Nuclear Industry, IAEA/SM- 172/42 (1973).

6.13 Gammon, J. R., Determination of feeding rates in Trichoptera using IN Radiation Emlogy Section, Health Physics Division,“CO, Annual Progress Report, ORNL-4446 (1970).

6.14 Harvey, R. S., Uptake and l~ssof radionuclides by the fresh-

water clam Lampsilis radiataI (gmel.), Health Phys. 17, 149-154 6.15\ilssS)* Harvey, R. S., Effects of temperature on the sorption of ,/radionuclides by a blue-green alga, Proc. Second National \‘ - Symposium on Radioecology, D. J. Nelson and F. C. Evans (eds. ) , CONF-676503 (1969). 1 6.16,\ Harvey, R. S., Temperature effects on the sorption of radio- \ -___.,’ nuclides by freshwater algae, Health Phys. -19, 293 (1970). 6.17 Kevern, N. R., and N. A. Griffith, Effect of trophic level on radionuclide accumulation by fish, IN Radiation Ecology Section, Health Physics Division, Annual Progress Report, ORNL-4007 (1966).

6.18 Lentsch, J. W., N. E. Wrenn, T. S. Kneip, and M. Eisenbud, Manmade radionuclides in the Hudson River Estuary, Fifth Ann. Health Phys. SOC. Midyear Topical Idaho Falls, Idaho (1970). Symposium, 6.19 Lowman, F. G., Iron and cobalt in ecology, IN Proc. First National Symposium in Radioecology, V. Schultz and A. W. Klement, 3r. ( eds. ) , Reinhold Pub1 . Corp., New York (1963). References for Section 3.6 (continued)

6.20 Lucas, H. F., Jr. and D, N. Edyi gtan, Concentrati s of trace elements in Great Lakes fishes, . Fish. Res. Bd. Can. -27, 677 (1970).

6.21 Marshall , J * S. and J. H. LeXoy, Iron, manganese, cobalt, and zinc cycles in a South Carolina reservoir, IN Proc. Third National Symposium QTI Radioecology, J. Nelson (ed. CONF-710501 (1973). D,

6.22 Mathis, R. J. and F. Cummings, Selected metals in sediments, water, and biota inT. the Illinois Rive?, J. Water Poll. Contr. Fed. 45, 1573 (1373). 6.23 Merlini , Margaret, Carla Big1 iocca A. Berg, and G. Pozzi , Trends in the concentration of heavy metals in organisms of a mesotrophic lake as determined by activation analysis , IN Broc. Symp. on Use of Nuclear Techniques in the Measurement and Control of Environmental Pollution, IAEA/SM-’142a/27 (1971).

6.24 Nerlini, Margaret, F. Girardi, and G. Pomzi, Activation analysis in studies of- an aquatic ecosystem, IN Proc. Syiip. on Nuclear Activation Techniques in the Life Sciences, IAEA/SM-91/7 (1967).

6.25 Morton, R. J., Status report no. 5 on Clinch River study, ORNIL-3721 Oak Ridge, Tennessee (1965).

6.26 Nelson, D. J., N. A. Griffith, J. bf. GQOC~,and S. A. Rucker, White Oak Lake Studies, IN Ecological Sciences Division Annual Report, July 31 1990, ORNL-4644 (197’1).

6.27 Nelson, D. J., N. A. Griffith, and S. A. Rucker, Radionuclide excretion studies, IN Radiation Ecology Section, Health Physics Division, Annual Progress Report, July 31 , 1964. ORNL-4446 (1 970) .

6.28 Nelson, D. J. and C, R. Malone, Feeding rates of snails determined with 60@o, IN Radiation Ecology Section, Health

Physics Division Annual Progress Report, July 31 .) 1958, ORNL-4316 (1969).

6.29 Opkel, I. L., and C, F. Fraser, The fate of cobalt-60 in a natural freshwater ecosystem, IN Proc. Third National Symposium on Radioecology, D. J. Nelson (ed. ) , CONF-710501 (1973). 177

References for Section 3.6 (continued)

6.30 Patten, B. C., N. A. Griff-ith, S. A. Rucker, and D. J. Nelson, Aquarium experiments, IN Radiation Ecology Section, Health Physics Division Annual Progress Report, July 31, 1968, ORNL-4316 (1969) .

6.31 Price, J. W., A study of the food habits of some Lake Erie fish, Bull. Ohio Biologics- Survey, New Series, Vol. 11, No. 1, Ohio State Univ., Columbus (1963).

6.32 Reed, J. R., Uptake and excretion by 6oCo by black bullheads Ictahnus melas (Rafinesque) , Health Phys. I21, 835 ( 1970). 6.33 Sherman, H. C., Essentials of nutrition, MacMillian Co., New York (1957).

6.34 Tirnofeev-Resovskii, N. V., E. A. Timofeeva-Resovskaya, G. A. Milyutinaya and A. E. Getsova, Coefficients of accumulation of radioactive isotopes of 16 different elements by freshwater organisms and the effect of EDTA on some of them, Doklady (Biol. Sci. Sect.) -132, 369 (1960). 6.35 Wiser, C. W. and D. J. Nelson, Uptake and elimination of 6oCo

by crayfish, Amer. Midl. Natur. I72, 181 (1964).

B IBL I[ OGRAPHY

181

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(1 958) I

APPE ND I X

27 3

EFFECT OF CARRIER CONCENTRATION ON THE DISCRIMINATION COEFFICIENT

The discrimination Coefficient, qi, is usually treated as a constant. The purpose of this appendix is to show that the coefficient can vary with concentration of the carrier in the water and in other sources. Treating i as a single compartment, the time rate of change in amount, Ri, of radionuclide R in organism or tissue i is:

d Ri -_ - Rik dt II * * and the steady-state amount, Ci , of carrier element C is given by

* - IT ci* - .- k* where

1,- = rate of input of radionuclide from all sources (UCilday) * IT = rate of input of carrier element from a91 sources (pg/day)

k = excretion rate coefficient for the radionuclide in i (day-'), * k = excretion rate coefficient for the carrier element in i at steady-state concentratio? (day-' 1. 21 4

At steady state, Eqs. (1-1) and (1-2) imply that

For the case where uptake is from a single prey j and from water:

I I -t- I and T = w j’ (1-4)

IT* = I W * 4 I.*J , 0-5) where

and I. are uptake rates (pCi/day) of a radionuclide from water Iw J and j, respectively; and and I are uptake rates (Uglday) of carrier Iw* j * element from water and j, respectively. For uptake from water:

Iw = a[RIw and,

Iw*= a*[C*IW where

a = uptzke rate coefficient for radionuclide from water (g/day), and

a* = uptake rate coefficient for carrier- element from water (g/day).

For uptake from food:

Ij = b[R],Q a.nd,

Ij* = b*[C*] J.O 21 5 where

b = absorbtion efficiency of radionuclide uptake from food (unitless),

b* = absorbtion efficiency of carrier element uptake from food (uni tless) , and

Q = feeding rate (g/day).

(1-10)

- k* --K (1-11)

Dividing Eq. (1-11) through by (R/C*)vl and incorporating the elimination coefficient of the prey, qj, gives

Equation (1-12) may be rewritten as

(1-13)

(1-14) 21 6

Because of homeostatic control , [C*lj is constant. Thus, as can

be seen from Equation (-1-141, qi should vary with [C*], since uptake from water is assumed proportional to [C*],. Equation (1-14) may be generalized for any set of soi~rces:

(1-15)

w 1-1 e re

rate of carrier elenient from j (pg/day), Iij* = intake by i and

aij = absorption efficiency of uptake of radionuclide from source j at surfaces of i (unitless), and

a = absorption efficiency of uptake of carrier element from iJ* source j at surfaces of i (unitless),

If water is a source, where j = 1 for water, ail is absorption of the

radionuclide brought to the body surfaces and q, E 1. 21 7

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78 r L. R. Anspaugh, University of California, Lawrence Livermore Laboratory, Eivermore , Cal iforni a 94550. 79. David S. Ballantine, Division of Biological and Environmental Research, U.S. Energy Research and Development Administration, Washington, D. C. 20545. 80. R. F. Barker, Product Standards Branch, Directorate of Regulatory Standards, U.S. Nuclear Regulatory Commission, Washington, D. C. 20545. 81. Nathaniel F. Barr, Division of Biomedical and Environmental Research, U.S. Energy Research and Development Administration, Washington, D. C. 20545. 82. D. S. Barth, National Environmental Research Center, EPA, P. 0 Box 15027, Las Vegas, Nevada 89114. 83. Richard t. Blanchard, Radiochemistry & Nuclear Engineering Facility, U.S. Environmental Protection Agency, Cincinnati, Ohio 45268. 84 * Andre Bouville, United Nations, Scientific Committee on the Effects of Atomic Radiation, New York, New York 10016. 85. J. A. Broadway, Eastern Environmental Research Facility, P. 0. Box 61, Montgomery, Alabama 36101, 86. Carter D, Broyles , Department Manager, Test Sciences (71 lo), Sandia Corp., P. 0. Box 5800, Albuquerque, New Mexico 87115. 21 8

87. W. W. Burr, Jr., Division of Biology and Medicine, U.S. Energy Research and Development Administration, Washington, 0. C. 20545. 88. M. W. Carter, Interdisciplinary Sciences Georgia Institute of Techno1 oyy , At1 anta, Georgia 20332. 89. Roger J. Clout’ier, Special Training Division, Oak Ridge Associated lJniversi ties, Oak Ridge, Tennessee 37830. 90. I?. G. Cochrell, Manager, Licensing Division, Westinghouse Electric Corporation, P. 0. Box 158, Madison, Pennsylvania

15663e 91 . Jerry Cohen, Lawrence Livermore Laboratory, P. 0. Box 808, Livermore, California 94551. 92. Raymond D. Cooper, Division of Bioinedical and Environmental Research, U. S. Energy Research and Devel opment Adriii ni stra- tion, Washington, D. C. 20545. 93. T. U. Crawford, Savannah River Laboratory, E. I. du Pont de NWI-IQU~Y & Company, Ai ken, South Carol i na 29801 . 94. R. G. Cuddihy, Inhalation Toxicology Research Institute, 5200 Gibson Blvd. SE, Albuquerque, New Mexico 87108. 95. Adrian H. Dahl, U.S. Energy Research and Development Administration, Idaho Operations Office, 550 2nd St., Idaho Falls, Idaho 83401. 96-98. Jared 3. Davis, Office of Nuclear Regulatory Research, Nuclear Regulatory Cornmi ssion, Washington, D. C. 20545.U.S. 99. 2. B. Dunaway, Office of Effects Evaluation, Nevada Operations Office, U.S. Energy Research and Development Administration, Las Vegas Nevada 87114. 100. G. G. Eichholz, School of Nuclear Engineering, Georgia Institute of Technology, Atlanta, Georgia 20332. 101. Rudolf J. Engleman, Division of Biomedical and Environmental Research, U. S. Energy Research and Development Adiiii nis tra ti on, Washington, D. C. 20545. 102. Edward H. Fleming, University of California Lawrence Livermore Laboratory, Mail Stop, L-1, Box 808 Livermore, Californl’a 94550. 103. R. Franklin, Division of Biomedical and Environmental Research, U.S. Energy Research and Development Administration, Washington, D. C. 20545. 104. Carl Gamertsfelder, Directorate of Licensing, U.S. Nuclear Regulatory Cormnission, Washington, D. C. 20545. 105. F. A. Gifford, Jr., Air Resources Atmospheric Turbulence and Diffusion Laboratory, NOAA, Oak Ridge, Tennessee 37830. 106. Abraham S. Goldin, Kresge Center for Environmental Health, Harvard Uni versi ty , School of Pub1 i c Heal th, Boston Massachusetts 021 00. 107. R. S. Goor, National Sci ence Foundati on, 21 01 Consti tuti on Ave., Washington, D. C. 20418. 108. R. L. Gotchy, Directorate of Licensing, U.S. Energy Research and Developiiient Administration, Washington, D. C. 20545. 109. Douglas Grahn, Argonne National Laboratory, 9700 South Cass Avenue, Argonne, Illinois 50439. 110. Harold J. Groh, Jr., Separstions Chemistry, Savannah River Laboratory, E. I. duPont d? Nemours and Company, P. 0. Box "A", Aiken , South Carol i na 29802. 111. D. F. Harmon, Office of Regulation, U.S. Nuclear Regulatory Commission, Bethesda, Maryland 20014. 112. Monte Hawkins, Head, Safety and Environmental Control, Allied General Nuclear Services, P. 0. Box 847 , Barnwell, South Carol i na 2981 2. 113. J. W. Healy, Health Physics Division, tos Alamos Scientific Laboratory, Box 1663, Los hlamos, New Mexico 87544. 174. John Hesl ep , Chief , Environmental Safety and Consumer Protec- tion Program, State Departnent of Public Health, 2151 Berkeley Way, Berkeley, California 97404. 115. George Hinman, Gulf Central Atomic, Box 608, San Diego,

Cal i forni a 921'1 2 a 116. I. A. Hobbs, Director, Technical and Production Division, Savannah River Operations Office, P. 0. Box "A", Aiken, South Carolina 29802. 117. John B. Hursh, Department of Radiation Biology and Biophysics, University of Rochester, Rochester, New York 14620. 118. Stephen Jinks, Institute of Environmental Medicine, Mew York University Medical Center, Long Meadow Road, Tuxedo, New York 10987. 119. B. Kahn, Interdisciplinary Sciences, Georgia Institute of Techno1 ogy , At1 anta, Georgia 20332. 120-1 69. Jacob Kastner, Radiological Assessment Branch, Directorate of Licensing, Nuclear Regulatory Commission, Washington, D. C. 20545. U.S. 170-1 71 . 3. R. Kercher , University of Cal i forni a, Lawrence Livemiore Laboratory, Livemore, California 94550. 172. 3. Kline, Argonne National Laboratory, Radiological Physics Division Ecological Studies, Argonne, Illinois 60439. 173. J. Koranda, Biomedical Division, Lawrence Radiation Laboratory, P. 0. Box 808 (L-13), Livemore, California 94551. 174. Paul Kruger, Professor of Nuclear Chemistry, Department of Civil Engineering, Stanford University, Stanford, California. 175. William Martin, Battelle Memorial Institute, 505 King Avenue, Columbus, Ohio 43201 . 176. P. J. Mellinger, EXXON Nuclear Company, 2955 George Washington Way, Richland, Washington 99352. 177. W. A. Mills, U.S. Environmental Protection Agency, Criteria and Standards Division, 401 M St., SW, Washington, D. C. 20460. 178. A. A. Moghissi, Office of Interdisciplinary Studies, Georgia Institute of Technology, Atlanta, Georgia 20332. 179. Daniel Iv(. Montgomery, Radiochemistry and Nuclear Engineering Faci 1 i ty , U .S . Envi ronmen tal Protection Agency, Ci nci nna ti , Ohio 45268. 180. K. Z. Morgan, School of Nuclear Engineering, Georgia Institute of Technology, Atlanta, Georgia 20332. 181. Henry W. Morton, Nuclear Fuel Services, Inc., 6000 Executive Blvd., Suite 600, Rockville, Maryland 20852.

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182. 1). R. Ne'son, Off-iez 3-f Rad-i:a"Lion Prog~arns,U.S. Environmental Pr-akection Age~y,401 1: Stx-pet, S. Washington, D. e. 20460. N., 183. Randolph R. Nw't~n,Division of Biological and Envfronmenta? Reseaxh, U.S. Energy Research and Developrrtent Administration,

Washington , D. C II 20545. 184. T. K. Noonan, UT-AEC Agricultural Research Laboratory, 1299 Betihr@l Valley Road, Oak Ridgt?, 'i-ennnssec 32830, 185. H. T. Odum , Department of Envi roivmental Engi nevi ng , Uni versi ty sL; Florida, Gainesville, Florida 32501. 186. J. S. Olson, Division of B5omedical and Environmental Reseilrch, U.S. Energy Research and Development fi,dm.irtistration, \dashington, 0. c. 20545. 187. W. S. Osburn, Divisiai-i of S4omredical and Environmental Research, U.S. Energy Reszarch and Dwclaprnent Admfnistration, Washington, D. c. 20545. 188. Charles Osterberg, of Biomedical and Environmental Research, U.S. EnergyDivision Rcseai-zh and De~elopmentAdministration, Washington, D. C. 20545. 189. Aat-on Padgett, Carolina Power atid Light, 1'. 0. BOX lStS'F, Raleiigh, North Carolina 27502. 130. C. C. Palmiter., Office of Radiation program^ Environmntal Proteec-tjon Agency, 401 M Street, SW, bhshingto~?,D. 2. 20460. 191. F. L Parker, Csepart~ient of Civi 1 Engi sleeri rig , Vande~bi1 t. Universi ty , M vi 11e , Tennessee 37203 192. Michael A. Par~mt,Radfolagical Assessment Branch, Directorate of Licensing, U. S. Nuclear Reg~latory@omission , Nastii ngton , D. C. 20545. 193. I!. R. Payne, 1J.S. Environmental Protection Agency Regaim IV, 1421 Peachtree St. , Atlanta, Georgia 303119. 194.. W. H. Penninghon, Divisioo OF B~ommedfcal and Environmental Research , U .S. Energy Research and kvel opment Wdiiii; nistrati on , !dashingtan, 10. 2. 20545. 195. J. F. Phillip, Directo?, Special Projects Division, San Franci sco Opera tinns Off?ce U . S . Energy Research and Devel oprnent Adriii n is trati on , 1 333 Brsa rhay , Oak 1and , Ca 1 i f orn ia 946 12. 196. K. P!~rushotha:fi~n, University of Missouri ai 691 la , Roll a Missouri 65401. 197. W. H. Ray, OPfiie of Re~~~la-tion,U.S. bhrlear Regulatory Corimission, Bethesda, Maryland 26014. 198 Ronald L. Ri l;si:hard, Unfversi ty sf Cali fornia , Lawrence Livermore babwatory , Livemore Ca9 ifemli.^ 34550. 199. I.. I_ RoSerts , Schaoll of Nucl ear Engi n~eii-1ng , Georgi a Ins ti tute of Technology, Atlai-ita, Gmirgia 20332.

200 I J. N. Rogers, Division 8327, Sandia Laboratories, P. 0. Box 969, Livemm-e, California 94559- 201. H. C. Rathschild, Radiation Protection Buwau, Brookfield Roads Ottawa K1 A1 C1 , Ontario , Canada. 202 I Keith J. Schiager, University of Plttsburg, Radlatltan Center, Pittsburg, Pennsylvania 15213. 293 e G. A. Selniel, Division of Riolo9y and MedSclne, Battelle Paci-ffe Northwst Laboratories, Rich;and, Washingtan 99352, 204. Seymour Shwi I Ier, Staff Glmber Joi tit Comi ttee on Atomic

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243. Research and Technical Support ERDA-OR0 244-439. distribution as shown in Division,under category for Given SciencesTID-4500 NTIS) Environmental and Earth (25 copies -

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